Environ. Sci. Technol. 2010, 44, 3468–3473
Biodegradation and Adsorption of Antibiotics in the Activated Sludge Process BING LI AND TONG ZHANG* Environmental Biotechnology Laboratory, Department of Civil Engineering, The University of Hong Kong, Hong Kong SAR, China
Received November 20, 2009. Revised manuscript received March 15, 2010. Accepted March 26, 2010.
The removal of 11 antibiotics of 6 classes, that is, two β-lactams (ampicillin and cefalexin), two sulfonamides (sulfamethoxazole and sulfadiazine), three fluoroquinolones (norfloxacin, ofloxacin, and ciprofloxacin), one tetracyclines (tetracycline), two macorlides (roxithromycin and anhydroerythromycin), and one others (trimethoprim), in activated sludge process was investigated using two series of batch reactors treating freshwater and saline sewage respectively. At environmental relevant concentrations tested in this study, biodegradation and adsorption were the major removal routes for the target antibiotics, where volatilization and hydrolysis were neglectable. Among the 11 target antibiotics, cefalexin and the two sulfonamides were predominantly removed by biodegradation in both freshwater and saline sewage systems. Ampicillin, norfloxacin, ciprofloxacin, ofloxacin, tetracycline, roxithromycin, and trimethoprim were mainly removed by adsorption. Divalent cations (Ca2+ and Mg2+) in saline sewage significantly decreased the adsorption of the three fluoroquinolones onto activated sludge. These three fluoroquinolones also exhibited certain biodegradability in the saline activated sludge reactor. Erythromycin-H2O was persistent in both saline and freshwater systems under the experimental conditions and could not be removed at all. Kinetics study showed that biodegradation of cefalexin, the two sulfonamides and the three fluoroquinolones followed first-order model well (R2: 0.921-0.997) with the rate constants ranging from 5.2 × 10-3 to 3.6 × 10-1 h-1.
Introduction The occurrence and fate of antibiotics in environment, including surface water, groundwater, and soils has drawn great attention of researchers all over the world in recent years (1, 2). Although the concentration of antibiotics residue in the environment is low, usually at ng/L to µg/L level in natural water (3) and wastewater (4), and µg/kg to mg/kg level in soil (5) and sludge (6), they are considered to be emerging pollutants because antibiotics and their transformation products, may result in the development/maintenance/transfer/spread of antibiotics resistant bacteria and antibiotics resistance genes in the long term and have serious impacts on the ecosystem (7–10). Antibiotics from three major sources, that is, industry, hospital, and household, are discharged into sewage (with or without in situ pretreatment) and then get into the * Corresponding author phone: +852-28578551; fax: +85225595337; e-mail:
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municipal wastewater treatment plants (WWTPs) (11). However, many studies have shown that the elimination of antibiotics in conventional WWTPs is not complete (8, 12–17) and they will finally enter into the environment via effluent or sludge. Therefore, WWTPs are one of the dominant point pollution sources for antibiotics (18). To date, most studies on antibiotics in WWTPs focused on their occurrences and concentrations in influent, effluent and sludge. The removal efficiencies (RE) of the detected antibiotics were usually calculated on the basis of the field sampling results. However, RE values obtained using the above strategy varied greatly among different studies, for example, from 22.7 (18) to 85.3% (4) for norfloxacin, from 14.6 (17) to 79.1% for erythromycin-H2O (19), and from 11.6 (20) to 85.4% (21) for tetracycline. Sometimes the RE values were even negative (18, 22–24). Additionally, although some antibiotics were found to be eliminated to some degree by the wastewater treatment process as shown above, the removal routes (biodegradation, adsorption, volatilization, or hydrolysis) are still not very clear. To overcome the uncertainty using field sampling at WWTPs and investigate the removal routes, a few pioneer works with well-controlled laboratory reactors have been conducted for several antibiotics at environmental relevant concentrations in activated sludge process (25–30). Pe´rez et al. (27) investigated the fate of three sulfonamides, that is, sulfamethazine, sulfamethoxazole, and sulfathiazole, at low concentration level (20 µg/L) in activated sludge process and found that about 50, 75, and 93% were removed via biodegradation in 10 days for the three antibiotics, respectively. Kim et al. (26) utilized two lab-scale sequencing batch reactors to study the removal of tetracycline (initial concentration of 250 µg/L) in activated sludge process. The results suggested that adsorption was the primary mechanism for tetracycline removal. Overall, up to now, the information on the removal and fate of antibiotics in WWTPs is still limited. The situation in Hong Kong is much more complicated and unique as some WWTPs (e.g., Shatin WWTP) treat saline sewage resulting from the practice of seawater toilet flushing. Thus, the aims of this study were (1) to measure the removal efficiency of 11 antibiotics of six different classes using two types of activated sludge treating freshwater sewage and saline sewage, respectively; (2) to investigate the elimination routes (biodegradation, adsorption, volatilization, and hydrolysis) for these antibiotics; (3) to study the biodegradation kinetics for the target antibiotics.
Experimental Section Target Antibiotics. According to the results of our previous study (22), totally, 11 antibiotics belonging to six different classes were detected in Hong Kong sewage, that is, two β-lactams: cefalexin (CLX) and ampicillin (AMP); two sulfonamides: sulfamethoxazole (SMX) and sulfadiazine (SDZ); three fluoroquinolones: norfloxacin (NOR), ofloxacin (OFL), and ciprofloxacin (CIP); one tetracyclines: tetracycline (TC); two macrolides: roxithromycin (ROX) and anhydro-erythromycin (ERY-H2O), and one others: trimethoprim (TMP). Consequently, these 11 antibiotics were selected as the target compounds in this study. Batch Experiment Design. Two series of batch experiments were conducted with fresh mixed liquor of aeration tanks sampled from two local wastewater plants, that is, Shatin WWTP treating saline sewage and Stanley WWTP treating freshwater sewage, with hydraulic retention times (HRT) of 10 and 17 h and sludge retention times (SRT) of 12 10.1021/es903490h
2010 American Chemical Society
Published on Web 04/12/2010
TABLE 1. Batch Experimental Design treatment
reactor no.
activated sludge
wastewater
antibiotics (100 µg/L)
caffeine (100 µg/L)a
0.1%b NaN3
aeration
removal routes
I II III IV controld
R1 and R1′ R2 and R2′ R3 R4 R5
+e + -f +
+ + + + +
+ + + + -
+ + + + +
+ + + -
+ + + +
B+A+V+Hc A+V+H V+H H -
a Caffeine was used as the reference organic chemical to indicate biological activity of sludge (Kim et al., 2008) as caffeine is known to be readily biodegradable and has no measurable adsorption to sludge. b NaN3 was used to inhibit the sludge biodegradation activity. c B-biodegradation; A-adsorption; V-volatilization; H-hydrolysis. d Control was used to monitor the effect of antibiotics on the sludge biological activity. e “+” indicated “with” or “presence”. f “-” indicated “without” or “absence”.
and 7 days, respectively. Both of the two WWTPs adopted anoxic-aerobic (A/O) activated sludge process with the average daily flows of 2.3 × 105 m3 and 8.8 × 103 m3, respectively. In the batch test, seven 2 L glass beakers with 1 L mixed liquor were run simultaneously at 22 ( 1 °C for 48 h following the five treatments (I, II, III, IV, and Control) shown in Table 1. The removal routes for antibiotics in activated sludge process are considered to be biodegradation (B), adsorption (A), volatilization (V), and hydrolysis (H) (27, 31). For Treatment I, R1 and R1′ were duplicate reactors in which all four removal routes occurred. For Treatment II, in duplicate reactors R2 and R2′, biodegradation was excluded because the activated sludge was inhibited by NaN3 (purity g99.5%, BDH, Limited Poole England). In Treatment III (R3), only volatilization and hydrolysis accounted for the elimination of antibiotics. Treatment IV (R4) and Control (R5) were utilized to investigate the hydrolysis of antibiotics and the effect of antibiotics on the sludge biological activity, respectively. According to the experimental results, the parts removed by biodegradation, adsorption and volatilization can be calculated based on differences between treatments as follows:
FIGURE 1. Removal of CLX in freshwater (FSS) and saline sewage systems (SSS). (eq 4), and second-order (SI eq S2) were applied to fit the biodegradation data. dC ) -k1·C S Ct ) C0·e-k1·t dt
(4)
Biodegradation ) I - II
(1)
Adsorption ) II - III
(2)
Where, C0 is initial concentration of the antibiotic; Ct is concentration of the antibiotic at time t; and k1 is the firstorder rate constant. Using this equation, half-lives, t1/2 can be calculated as (ln 2)/(k1).
Volatilization ) III - IV
(3)
Results and Discussion
All reactors were placed in a closed chamber to avoid possible photolysis. The antibiotics were spiked with an aqueous stock solution mixture containing 15 mg/L of each antibiotic to obtain a final concentration of 100 µg/L. The aeration and mixing were supplied by the magnetic stirrers at 150 rpm. Samples of the slurry were taken from the batch reactors at the following times: 0, 0.25, 0.5, 1, 2, 5, 10, 15, 24, 36, 48 h. Other operational parameters were summarized in Supporting Information (SI) Table S2. Sample Analysis via UPLC/MS/MS. Slurry samples of 5 mL were withdrawn from reactors using glass syringe and filtered via 0.2 µm cellulose nitrate membrane. The first 2 mL filtrate was discarded and the following 1.5 mL was collected in an amber vial. All the samples were kept in dark at 4 °C and analyzed directly via Acquity ultraperformance liquid chromatographystandem mass spectrometry (UPLCMS/MS, Waters) within 24 h. UPLC-MS/MS was operated in the positive electrospray ionization multiple reaction monitoring (MRM) mode. Other detailed information, such as mobile phase, column temperature, flow rate, MRM parameters, antibiotic standards, formic acid, and cellulose nitrate membrane were reported in the previous study (22). Biodegradation Kinetics Models. Three biodegradation kinetics models, that is, zero-order (SI eq S1), first-order
Removal of Antibiotics in Activated Sludge. In this study, all target antibiotics were found to be stable and no hydrolysis occurred during the testing period (48 h) and their removal due to volatilization can be ignored on the basis of the data of Treatment III and IV. The absence of volatilization was also in agreement with their ultra low Henry’s Law constants (4.97 × 10-31 ∼ 1.58 × 10-10, SI Table S1), greater molecular weight (>200, SI Table S1) as well as the presence of several polar groups in the target antibiotics. Thus, only biodegradation and adsorption were discussed in details in the following sections. Comparison of the caffeine biodegradation profile in Treatment Control with that in Treatment I (SI Figure S1) suggested that the impact of antibiotics on the sludge biological activity was neglectable. In addition, no caffeine was degraded in Treatment II in 48 h while it was completely biodegraded within 10 h in Treatment I, indicating that the biological activity of sludge in Treatment II was completely inhibited by NaN3 (SI Figure S2). β-Lactams. As shown in Figure 1, CLX was significantly (up to 97.3%) removed via biodegradation without lag phase in both freshwater sewage system (FSS) and saline sewage system (SSS) in 10 h. A widely accepted interpretation is that β-lactam ring is unstable and can be cleaved by β-lactamases, a group of widespread enzymes in bacteria (32). However, the fate of AMP, another β-lactams, in the activated sludge VOL. 44, NO. 9, 2010 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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FIGURE 2. Removal of SDZ in freshwater (FSS) and saline sewage systems (SSS).
FIGURE 3. Removal of NOR in freshwater (FSS) and saline sewage systems (SSS).
process was totally different from CLX. As shown in SI Figure S3, the major removal route for AMP was adsorption rather than biodegradation. About 56.8 and 38.7% of AMP were adsorbed rapidly by the activated sludge during the first 15 min in FSS and SSS, respectively, and adsorption equilibrium was reached after 10 h for both of two systems. CLX and AMP belong to two subclasses of β-lactams, that is, cephalosporins and penicillins, respectively. It is not surprising that their removals in activated sludge process were pretty different. The limited studies on transformation and fate of β-lactams in wastewater treatment processes showed that the removal routes for other antibiotics in this class, including amoxicillin, benzylpenicillin, and ceftriaxone also varied greatly, possibly due to their different chemical structures, such as the variable side chains. Amoxicillin was removed by 90% through biodegradation and adsorption in 5 h in standardized batch tests using activated sludge, and the removal was dominated by biodegradation (33). Using 14 C-labeling method, it was found that about 25% of benzylpenicillin (28 µg/L) was mineralized, whereas ceftriaxone (14 µg/L) was not mineralized at all (30). Although CLX and AMP behaved differently in the activated sludge process, the removal mechanisms and removal efficiencies of each individual of them were quite similar in the freshwater and saline sewage systems. Sulfonamides. As shown in Figure 2 and SI Figure S4, removal of SDZ and SMX via adsorption was neglectable, being consistent with results of other researches showing that sulfonamides were highly water-soluble and had negligible adsorption to activated sludge (21, 25, 27). For SDZ, the removal efficiencies after 48 h in SSS and FSS were 37.3 and 53.4%, respectively. The similar results were also obtained for SMX, that is, 22.6 and 39.1% being eliminated via biodegradation in 48 h in saline and freshwater sewage, respectively. This implied the different sulfonamidesdegrading abilities of microbial communities in the two systems. The amount and activity of the sulfonamides degraders in the activated sludge deserve further quantification and characterization. Although SMX and SDZ exhibited significant biodegradation in the batch reactors run for 48 h, their removal efficiencies would be only 9.9∼17.1% (SMX) and 5.4∼23.8% (SDZ) in Shatin and Stanley WWTPs, respectively, considering HRT values applied in the two WWTPs, that is, 10 h and 17 h. Therefore, the elimination of sulfonamides in these two WWTPs by activated sludge treatment is insignificant. Fluoroquinolones. NOR, OFL, and CIP were eliminated quickly by adsorption in the first 15 min in both FSS and SSS (Figure 3, SI Figures S5 and S6). In FSS, adsorption was the dominant removal route for the three fluoroquinolones, and the corresponding removal efficiencies were 91.6, 84.4, and 90.8%, respectively. Similar results were reported by Batt et al. (21) that CIP was adsorbed significantly (removal efficiency
of ∼85%) and no biodegradation happened in 6 h. In SSS, both adsorption and biodegradation played important roles in the removal of the target fluoroquinolones. In the first 15 min, the removal was dominated by adsorption and adsorption equilibrium was reached rapidly, resulting in the adsorption elimination efficiencies of 60.5% (NOR), 42.3% (OFL), and 52.8% (CIP). After that, biodegradation began to happen and achieved removal efficiencies of 26.7% (NOR), 40.8% (OFL), and 32.2% (CIP) at the end of the 48 h experiment. In fact, although the three fluoroquionolones could be biodegraded to some degree in saline sewage system, it should be noted that they cannot be removed significantly via biodegradation considering the short HRT (10 h) applied in Shatin WWTP. Judging from the adsorption efficiency, it can be concluded that the adsorption capability of freshwater activated sludge for these three antibiotics were significantly higher than that of saline activated sludge. This might be due to the effect of divalent cations in the saline sewage as fluoroquinlones are able to form stable complexes with divalent metalions,includingCa2+,Mg2+,etc(34,35).Theoctanol-water partition coefficients (Kow) can be utilized to assess the hydrophobic adsorption for uncharged molecules (6). In this way, uncharged antibiotics with pKow < 2.5 are considered to have low adsorption potential (36). However, NOR, OFL, and CIP exhibited high adsorption potential onto activated sludge (367.2 L/kg < Kd < 664.8 L/kg in SSS and 2256.9 L/kg < Kd < 5122.7 L/kg in FSS) despite their low pKow values (-1.03∼0.28, see Table S1). Considering their zwitterionic characteristics (pKaCOOH ) 3.01∼6.26, pKaNH2 ) 7.65∼10.58), it could be concluded that the adsorption of fluoroquinolones on activated sludge was mainly affected by the electrostatic force rather than hydrophobic interaction (6). Another interesting phenomenon is that all the three fluoroquinolones were partially biodegraded in SSS while no biodegradation was observed in FSS. This might be mainly due to the different microbial communities in the two activated sludges and some special fluoroquinolonesdegrading microorganisms bacteria might only exist in the saline activated sludge. Tetracyclines. As shown in Figure 4, TC was adsorbed significantly and rapidly onto activated sludge with no biodegradation, being consistent with a few earlier studies (21, 26). The adsorption rate was up to ∼90% during the first 15 min in FSS and SSS. At the adsorption equilibrium, the removal efficiencies in the two systems were 98.0 and 92.3%, respectively. Unlike the adsorption of fluoroquinolones which was significantly different in two systems, it seems that seawater components in the saline sewage had minor impact on TC adsorption. This might be due to the following three reasons. First, the effect of ionic strength on TC adsorption was insignificant at the low surface coverage (0.08-0.1 mmol/
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TABLE 2. First-Order Kinetics Model Parameters first-order kinetics: Ct ) C0 × eantibiotic reactor system C0 (µg/L)
FIGURE 4. Removal of TC in freshwater (FSS) and saline sewage systems (SSS). kg) applied in this study (37). Second, with pKa values of 3.3, 7.7, and 9.7 (38), TC molecule may exist as the cation (TC+00), zwitterion (TC+-0), monoanion (TC+--) and dianion (TC0--) form, respectively. At the pH values (6.9-7.3) applied in this study, the predominant forms should be zwitterion and monoanion species with the ratio from about 85:15 to 70:30 (39), dominated by zwitterion. It was reported that zwitterion adsorption was relatively insensitive to ionic strength effects (38, 40). Third, the complexation of TC with divalent metal ions was neglectable since complexation occurs predominantly under basic condition (41). It was reported that only at pH g 9, which is far beyond the pH range in the study, the complexation of TC with Ca2+ occurred significantly (42). Macrolides. As shown in Figure S7 and Figure S8, the removal of the two macrolides antibiotics, that is, ROX and ERY-H2O, in activated sludge process were quite different. For ROX, it was removed mainly via adsorption while no biodegradation occurred. The removal efficiencies after 48 h were 34.6 and 30.9% in FSS and SSS, slightly lower than the results (>50%) reported previously (43). For ERY-H2O, no removal was observed in both two systems, indicating that it was persistent and could not be removed in the aerobic biological treatment process. Trimethoprim. TMP exhibited certain adsorptivity to activated sludge and no biodegradability (SI Figure S9). The removal efficiency via adsorption at the end of the experiment was 26.4 and 19.0% in FSS and SSS, respectively. This was in agreement with the results of a few studies which showed that TMP was eliminated by ∼10% via adsorption (21) and no biodegradation occurred in conventional activated sludge process (27). However, some researchers found that TMP could be biodegraded effectively by nitrifying activated sludge (27, 28, 44). According to Batt et al. (28), when the activity of the nitrifying bacteria was inhibited, the removal percentage of TMP decreased from 70 to 25% in batch reactors. The results of their batch test were also corroborated by the data in a full-scale municipal WWTP. Removal efficiency of TMP was up to 50% for the nitrifying activated sludge, whereas only about 1% for the conventional activated sludge. Considering the short SRT values (7 and 12 days, respectively) applied in the activated sludge processes of Stanley and Shatin, it is reasonable to assume that there is no biodegradation of TMP in these two WWTPs. Although SI Figure S1 indicated no impact of the target antibiotics on the biodegradation kinetics of caffeine, it should be noted that caffeine, as a readily biodegradable substrate for activated sludge bacteria, is different from those tested antibiotics displaying partial biodegradation, that is, SDZ, SMX, NOR, OFL, and CIP. Although theoretically the antibiotics with initial total concentration of 1100 µg/L might compromise the biodegradation of antibiotics, this simulated the real situation which contains a large number of different
-1
k1 (h )
k1 · t
R2
t1/2 (h)
CLX
FSS SSS
104.7 92.9
2.6 × 10-1 3.6 × 10-1
SMX
FSS SSS
93.4 104.0
64.2 0.992 1.1 × 10-2 5.2 × 10-3 133.3 0.921
SDZ
FSS SSS
95.3 101.5
1.6 × 10-2 8.6 × 10-3
NOR
SSS
95.8
6.2 × 10-3 111.8 0.936
CIP
SSS
96.7
7.5 × 10-3
92.4 0.923
99.0
-2
66.6 0.972
OFL
SSS
1.0 × 10
2.6 0.994 1.9 0.997
42.0 0.995 80.6 0.968
antibiotics in the municipal wastewater and the total concentration might be at sub-mg/L to mg/L levels. Biodegradation Kinetics of Antibiotics. It is well accepted that biodegradation of trace amount organic in sewage follows the first-order kinetics model well in biological treatment due to the relatively low substrate concentration compared to the biomass concentrations (45). As expected, compared with zero-order and second-order model, the biodegradation of antibiotics fitted first-order kinetics model well (Table 2 and SI Table S3) at the low concentration of 100 µg/L applied in this study. β-Lactams. The biodegradation kinetics of CLX followed the first-order model (SI Figure S10) perfectly (R2 > 0.994) and the corresponding parameters including C0, k1, and t1/2 were summarized in Table 2. Judging from the short halflives (2.6 and 1.9 h) in two systems, it is concluded that CLX was easily biodegraded and would be effectively removed from the sewages within the designed HRT values (Stanley: 17 h and Shatin: 10 h). The values of biodegradation rate constant (k1), 0.26 h-1 in FSS and 0.36 h-1 in SSS, indicated that the biodegradation of CLX in the latter system was slightly faster than the former one. Sulfonamides. The biodegradation kinetics of two sulfonamides antibiotics followed the first-order model well with all R2 values ranging from 0.921-0.995 (SI Figures S11 and S12). Other fitting parameters were summarized in Table 2. The half-lives of SMX and SDZ were so long (42.0-133.3 h) that the biodegradation of sulfonamides within the designed HRTs in Stanley and Shatin WWTPs was neglectable. On the basis of the first-order rate constants, k1, it can be concluded that the biodegradation rate of SDZ was faster than that of SMX. In addition, for the same sulfonamides, the biodegradation rate in FSS was higher than that in SSS. Fluoroquinolones. The biodegradation of fluoroquinolones also fitted the first-order model well with all R2 values ranging from 0.923-0.972 (SI Figure S13 and Table 2). The half-lives for the three fluoroquinolones were up to 66.6-111.8 h, suggesting that the biodegradation is insignificant in WWTPs. Judging from k1 values of these three antibiotics, the biodegradation rate followed the order of OFL > CIP > NOR. The first-order rate constant,k1, is an apparent constant to indicate the biodegradation rate of antibiotics in FSS and SSS. To compare the antibiotics biodegradation activity of activated sludge in different treatment systems, specific firstorder rate constant, k1,specific, which was normalized by the MLSS was used in this study (SI Table S4). Due to the limited information, only k1,specific values of SDZ (1.29 × 10-1 L · gMLSS-1 · d-1) and SMX (1.90 × 10-1 L · gMLSS-1 · d-1) in FSS were reported by Ingerslev et al. (25) and Abegglen et al. (43). Compared with the k1,specific values of 1.74 × 10-1 and 1.20 × VOL. 44, NO. 9, 2010 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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10-1 L · gMLSS-1 · d-1 for SDZ and SMX in the present study, it can be concluded that the sulfonamides biodegradation activity of activated sludge in FSS was comparable with those reported in the above references. In SSS, the k1,specific values for SDZ and SMX were 8.02 × 10-2 and 4.85 × 10-2 L · gMLSS-1 · d-1, respectively. According to the k1,specific values in FSS and SSS, it was found that the activated sludge in FSS was more effective in degrading SDZ and SMX while less effective in degrading CLX, NOR, CIP and OFL than the activated sludge in SSS. k1,specific can also be used as a parameter to compare the biodegradability of different antibiotics in the same treatment system. In FSS, the biodegradability followed the order of CLX.SDZ > SMX and no biodegradation occurred for the three fluoroquinolones. In SSS, the biodegradability followed the order of CLX.OFL > SDZ > CIP > NOR > SMX. The k1,specific values of CLX was much greater (more than an order of magnitude) than those of the other antibiotics in both FSS and SSS. According to the classification system suggested by Abegglen et al. (43), CLX is a readily biodegradable substance while SDZ, SMX, OFL, CIP, and NOR are slowly biodegradable substance based on their k1,specific values. Furthermore, with given values of HRT and MLSS, the removal efficiencies of antibiotics in different activated sludge systems could be predicted based on the k1,specific values. For Shatin WWTP, with HRT of 10 h and MLSS of 2572 mg/L, the removal efficiencies via biodegradation for CLX, OFL, SDZ, CIP, NOR, and SMX were calculated as 97.3, 9.5, 8.2, 7.2, 6.0, and 5.1%, respectively. While for Stanley WWTP (HRT: 17 h; MLSS: 2210 mg/L), the removal efficiencies of CLX, SDZ, and SMX via biodegradation were calculated as 98.8, 23.8, and 17.1%, respectively.
Acknowledgments We thank the Hong Kong General Research Fund (HKU7202/ 09E) for the financial support of this study, and B. Li thanks The University of Hong Kong for the postgraduate studentship.
Supporting Information Available Figure S1-Figure S13 and Table S1-Table S4. This material is available free of charge via the Internet at http:// pubs.acs.org.
Literature Cited (1) Xiao, Y.; Chang, H.; Jia, A.; Hu, J. Y. Trace analysis of quinolone and fluoroquinolone antibiotics from wastewaters by liquid chromatography-electrospraytandemmassspectrometry.J.Chromatogr., A. 2008, 1214 (1-2), 100–108. (2) Ku ¨ mmerer, K. Pharmaceuticals in the Environment: Sources, Fate, Effects and Risks, 1st ed.; Springer-Verlag: Berlin, 2001. (3) Diaz-Cruz, M. S.; Garcia-Galan, M. J.; Barcelo, D. Highly sensitive simultaneous determination of sulfonamide antibiotics and one metabolite in environmental waters by liquid chromatographyquadrupole linear ion trap-mass spectrometry. J. Chromatogr., A. 2008, 1193 (1-2), 50–59. (4) Watkinson, A. J.; Murby, E. J.; Costanzo, S. D. Removal of antibiotics in conventional and advanced wastewater treatment: Implications for environmental discharge and wastewater recycling. Water Res. 2007, 41 (18), 4164–4176. (5) Hamscher, G.; Sczesny, S.; Hoper, H.; Nau, H. Determination of persistent tetracycline residues in soil fertilized with liquid manure by high-performance liquid chromatography with electrospray ionization tandem mass spectrometry. Anal. Chem. 2002, 74 (7), 1509–1518. (6) Golet, E. M.; Xifra, I.; Siegrist, H.; Alder, A. C.; Giger, W. Environmental exposure assessment of fluoroquinolone antibacterial agents from sewage to soil. Environ. Sci. Technol. 2003, 37 (15), 3243–3249. (7) Boxall, A. B. A.; Kolpin, D. W.; Halling-Sorensen, B.; Tolls, J. Are veterinary medicines causing environmental risks. Environ. Sci. Technol. 2003, 37 (15), 286A–294A. (8) Go¨bel, A.; Thomsen, A.; McArdell, C. S.; Joss, A.; Giger, W. Occurrence and sorption behavior of sulfonamides, macrolides, and trimethoprim in activated sludge treatment. Environ. Sci. Technol. 2005, 39 (11), 3981–3989. 3472
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(9) Hernando, M. D.; Mezcua, M.; Fernandez-Alba, A. R.; Barcelo, D. Environmental risk assessment of pharmaceutical residues in wastewater effluents, surface waters and sediments. In 1st Workshop on Validation of Robustness of Sensors and Bioassays for Screening Polluatants, Menorca, Spain, December 02-03, 2004; pp 334-342. (10) Martinez, J. L. Antibiotics and antibiotic resistance genes in natural environments. Science. 2008, 321 (5887), 365–367. (11) Giger, W.; Alder, A. C.; Golet, E. M.; Kohler, H. P. E.; McArdell, C. S.; Molnar, E.; Siegrist, H.; Suter, M. J. F. Occurrence and fate of antibiotics as trace contaminants in wastewaters, sewage sludges, and surface waters. Chimia. 2003, 57 (9), 485–491. (12) Golet, E. M.; Alder, A. C.; Giger, W. Environmental exposure and risk assessment of fluoroquinolone antibacterial agents in wastewater and river water of the Glatt Valley Watershed, Switzerland. Environ. Sci. Technol. 2002, 36 (17), 3645–3651. (13) Ku ¨ mmerer, K. Significance of antibiotics in the environment. J. Antimicrob. Chemother. 2003, 52 (1), 5–7. (14) McArdell, C. S.; Molnar, E.; Suter, M. J. F.; Giger, W. Occurrence and fate of macrolide antibiotics in wastewater treatment plants and in the Glatt Valley Watershed, Switzerland. Environ. Sci. Technol. 2003, 37 (24), 5479–5486. (15) Miao, X. S.; Bishay, F.; Chen, M.; Metcalfe, C. D. Occurrence of antimicrobials in the final effluents of wastewater treatment plants in Canada. Environ. Sci. Technol. 2004, 38 (13), 3533– 3541. (16) Go¨bel, A.; Thomsen, A.; McArdell, C. S.; Alder, A. C.; Giger, W.; Theiss, N.; Loffler, D.; Ternes, T. A. Extraction and determination of sulfonamides, macrolides, and trimethoprim in sewage sludge. J. Chromatogr., A. 2005, 1085 (2), 179–189. (17) Xu, W. H.; Zhang, G.; Li, X. D.; Zou, S. C.; Li, P.; Hu, Z. H.; Li, J. Occurrence and elimination of antibiotics at four sewage treatment plants in the Pearl River Delta (PRD), South China. Water Res. 2007, 41 (19), 4526–4534. (18) Gulkowska, A.; Leung, H. W.; So, M. K.; Taniyasu, S.; Yamashita, N.; Yeunq, L. W. Y.; Richardson, B. J.; Lei, A. P.; Giesy, J. P.; Lam, P. K. S. Removal of antibiotics from wastewater by sewage treatment facilities in Hong Kong and Shenzhen, China. Water Res. 2008, 42 (1-2), 395–403. (19) Karthikeyan, K. G.; Meyer, M. T. Occurrence of antibiotics in wastewater treatment facilities in Wisconsin, USA. Sci. Total Environ. 2006, 361 (1-3), 196–207. (20) Spongberg, A. L.; Witter, J. D. Pharmaceutical compounds in the wastewater process stream in Northwest Ohio. Sci. Total Environ. 2008, 397 (1-3), 148–157. (21) Batt, A. L.; Kim, S.; Aga, D. S. Comparison of the occurrence of antibiotics in four full-scale wastewater treatment plants with varying designs and operations. Chemosphere. 2007, 68 (3), 428– 435. (22) Li, B.; Zhang, T.; Xu, Z. Y.; Fang, H. H. P. Rapid analysis of 21 antibiotics of multiple classes in municipal wastewater using ultra performance liquid chromatography-tandem mass spectrometry. Anal. Chim. Acta. 2009, 645 (1-2), 64–72. (23) Lindberg, R. H.; Wennberg, P.; Johansson, M. I.; Tysklind, M.; Andersson, B. A. V. Screening of human antibiotic substances and determination of weekly mass flows in five sewage treatment plants in Sweden. Environ. Sci. Technol. 2005, 39 (10), 3421– 3429. (24) Roberts, P. H.; Thomas, K. V. The occurrence of selected pharmaceuticals in wastewater effluent and surface waters of the lower Tyne catchment. Sci. Total Environ. 2006, 356 (1-3), 143–153. (25) Ingerslev, F.; Halling-Sorensen, B. Biodegradability properties of sulfonamides in activated sludge. Environ. Toxicol. Chem. 2000, 19 (10), 2467–2473. (26) Kim, S.; Eichhorn, P.; Jensen, J. N.; Weber, A. S.; Aga, D. S. Removal of antibiotics in wastewater: Effect of hydraulic and solid retention times on the fate of tetracycline in the activated sludge process. Environ. Sci. Technol. 2005, 39 (15), 5816–5823. (27) Pe´rez, S.; Eichhorn, P.; Aga, D. S. Evaluating the biodegradability of sulfamethazine, sulfamethoxazole, sulfathiazole, and trimethoprim at different stages of sewage treatment. Environ. Toxicol. Chem. 2005, 24 (6), 1361–1367. (28) Batt, A. L.; Kim, S.; Aga, D. S. Enhanced biodegradation of iopromide and trimethoprim in nitrifying activated sludge. Environ. Sci. Technol. 2006, 40 (23), 7367–7373. (29) Joss, A.; Zabczynski, S.; Gobel, A.; Hoffmann, B.; Loffler, D.; McArdell, C. S.; Ternes, T. A.; Thomsen, A.; Siegrist, H. Biological degradation of pharmaceuticals in municipal wastewater treatment: Proposing a classification scheme. Water Res. 2006, 40 (8), 1686–1696.
(30) Junker, T.; Alexy, R.; Knacker, T.; Kummerer, K. Biodegradability of 14C-labeled antibiotics in a modified laboratory scale sewage treatment plant at environmentally relevant concentrations. Environ. Sci. Technol. 2006, 40 (1), 318–324. (31) Kim, S.; Aga, D. S. Potential ecological and human health impacts of antibiotics and antibiotic-resistant bacteria from wastewater treatment plants. J. Toxicol. Environ. Health, Part B. 2007, 10 (8), 559–573. (32) Cha, J. M.; Yang, S.; Carlson, K. H. Trace determination of β-lactam antibiotics in surface water and urban wastewater using liquid chromatography combined with electrospray tandem mass spectrometry. J. Chromatogr., A. 2006, 1115 (1-2), 46–57. (33) Andreozzi, R.; Caprio, V.; Ciniglia, C.; De Champdore, M.; Lo Giudice, R.; Marotta, R.; Zuccato, E. Antibiotics in the environment: Occurrence in Italian STPs, fate, and preliminary assessment on algal toxicity of amoxicillin. Environ. Sci. Technol. 2004, 38 (24), 6832–6838. (34) El-Kommos, M. E.; Saleh, G. A.; El-Gizawi, S. M.; Abou-Elwafa, A. Spectrofluorometric determination of certain quinolone antibacterials using metal chelation. Talanta. 2003, 60 (5), 1033– 1050. (35) Turiel, E.; Martin-Esteban, A.; Tadeo, J. L. Multiresidue analysis of quinolones and fluoroquinolones in soil by ultrasonic-assisted extraction in small columns and HPLC-UV. Anal. Chim. Acta. 2006, 562 (1), 30–35. (36) Rogers, H. R. Sources, behaviour and fate of organic contaminants during sewage treatment and in sewage sludges. Sci. Total Environ. 1996, 185 (1-3), 3–26. (37) Gu, C.; Karthikeyan, K. G. Interaction of tetracycline with aluminum and iron hydrous oxides. Environ. Sci. Technol. 2005, 39 (8), 2660–2667. (38) Figueroa, R. A.; Leonard, A.; Mackay, A. A. Modeling tetracycline antibiotic sorption to clays. Environ. Sci. Technol. 2004, 38 (2), 476–483.
(39) Chang, P. H.; Li, Z. H.; Yu, T. L.; Munkhbayer, S.; Kuo, T. H.; Hung, Y. C.; Jean, J. S.; Lin, K. H. Sorptive removal of tetracycline from water by palygorskite. J. Hazard. Mater. 2009, 165 (1-3), 148–155. (40) Parolo, M. E.; Savini, M. C.; Valles, J. M.; Baschini, M. T.; Avena, M. J. Tetracycline adsorption on montmorillonite: pH and ionic strength effects. Appl. Clay Sci. 2008, 40 (1-4), 179– 186. (41) Wessels, J. M.; Ford, W. E.; Szymczak, W.; Schneider, S. he complexation of tetracycline and anhydrotetracycline with Mg2+ and Ca2+: A spectroscopic study. J. Phys. Chem. B. 1998, 102 (46), 9323–9331. (42) Lambs, L.; Decocklereverend, B.; Kozlowski, H.; Berthon, G. Metal ion-tetracycline interactions in biological-fluids. 9. Circular-dichroism spectra of calcium and magnesium complexes with tetracycline, oxytetracycline, doxycycline, and chlortetracycline and discussion of their binding modes. Inorg. Chem. 1988, 27 (17), 3001–3012. (43) Abegglen, C.; Joss, A.; McArdell, C. S.; Fink, G.; Schlusener, M. P.; Ternes, T. A.; Siegrist, H. The fate of selected micropollutants in a single-house MBR. Water Res. 2009, 43 (7), 2036–2046. (44) Eichhorn, P.; Ferguson, P. L.; Perez, S.; Aga, D. S. Application of ion trap-MS with QqTOF-MS in the identification H/D exchange and of microbial degradates of trimethoprim in nitrifying activated sludge. Anal. Chem. 2005, 77 (13), 4176– 4184. (45) Metcalf; Eddy. Wastewater Engineering: Treatment and Reuse; McGraw Hill: Singapore, 2004.
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