Occurrence and removal of antibiotics in a municipal wastewater reclamation plant in Beijing, China

Chemosphere xxx (2013) xxx–xxx

Contents lists available at SciVerse ScienceDirect

Chemosphere journal homepage: www.elsevier.com/locate/chemosphere

Occurrence and removal of antibiotics in a municipal wastewater reclamation plant in Beijing, China Wenhui Li a, Yali Shi a,⇑, Lihong Gao a,b, Jiemin Liu b, Yaqi Cai a a b

State Key Laboratory of Environmental Chemistry and Ecotoxicology, Research Center for Eco-Environmental Science, Chinese Academy of Sciences, Beijing 100085, China School of Chemistry and Biological Engineering, University of Science and Technology Beijing, University of Science and Technology Beijing, Beijing 100083, China

h i g h l i g h t s " Antibiotics are ubiquitous in aqueous and sludge samples in WRP. " Most antibiotics are effectively removed through advance treatment. " Algae and aquatic plants may be at risk of antibiotics in reused water. " Organic carbon may influence distribution of antibiotics between water and sludge.

a r t i c l e

i n f o

Article history: Received 22 August 2012 Received in revised form 31 December 2012 Accepted 8 January 2013 Available online xxxx Keywords: Antibiotics Quinolones Sulfonamides Macrolides Wastewater reclamation plant

a b s t r a c t In this study, we investigated the occurrences and fates of eight quinolones (QNs), nine sulfonamides (SAs), and five macrolides (MCs) in a wastewater reclamation plant (WRP) in Beijing, China. Among all the 22 antibiotics considered, quinolones were the dominant antibiotics in all samples (4916 ng L1 in influents, 1869 ng L1 in secondary effluents, 123 ng L1 in tertiary effluents, and 9200 lg kg1 in sludge samples), followed by sulfonamides (2961 ng L1 in influents, 1053 ng L1 in secondary effluents, 25.9 ng L1 in tertiary effluents, and 63.7 lg kg1 in sludge samples) and macrolides (365 ng L1 in influents, 353 ng L1 in secondary effluents, 24.7 ng L1 in tertiary effluents, and 32.7 lg kg1 in sludge samples). The removal efficiencies of the target antibiotics were limited (32 to 78%) in the conventional treatment. This study indicated that quinolones were mainly removed from the secondary clarifier, and sulfonamides were degraded in the oxic tank; while macrolides were considered as persistent during the conventional treatment. After the advance treatment, the target antibiotics could be effectively removed at high rates (85–100%), and the risks of antibiotic contamination significantly decreased. However, risk assessment showed that the risk of ofloxacin and erythromycin on organisms in recycled water needed further investigations. Ó 2013 Elsevier Ltd. All rights reserved.

1. Introduction As an emerging group of environmental contaminant, antibiotics have attracted growing attention due to their potential undesirable effects on human health and aquatic ecosystems (Sebastine and Wakeman, 2003; Samanidou and Evaggelopoulou, 2007; Dirany et al., 2011; Yang et al., 2011). Antibiotics are widely used to treat infectious diseases for both humans and animals, and to promote the growth of animals in livestock farming, aquaculture and agriculture (Gao et al., 2012). However, many antibiotics cannot be completely absorbed or metabolized in the body, and most of them are introduced through excretion into the sewage

⇑ Corresponding author. Tel./fax: +86 10 62849182. E-mail address: [email protected] (Y. Shi).

system, which has become an important source of antibiotics in the environment (Tamtam et al., 2011; Leung et al., 2012). Municipal wastewater system is regarded as the major barrier for these antibiotics from wastewater into aquatic environment (Duong et al., 2008; Al-Rifai et al., 2011). However, previous studies have shown that many antibiotics are only partially eliminated in conventional wastewater treatment plants (WWTPs), with removal efficiencies ranging from 80 to 96% for most selected antibiotics (Clara et al., 2005; Onesios et al., 2009; Le-Minh et al., 2010). As a consequence, these antibiotics may be consistently present in effluents, and the concentrations of some of them are even higher in effluents than in influents (Lin et al., 2009). It has been reported that additional process, such as ozonation or membrane ultrafiltration, can reduce the concentrations of antibiotics in the effluents of WWTPs (Zorita et al., 2009). For example, Watkinson et al. reported microfiltration/reverse osmosis reduced the

0045-6535/$ - see front matter Ó 2013 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.chemosphere.2013.01.040

Please cite this article in press as: Li, W., et al. Occurrence and removal of antibiotics in a municipal wastewater reclamation plant in Beijing, China. Chemosphere (2013), http://dx.doi.org/10.1016/j.chemosphere.2013.01.040

2

W. Li et al. / Chemosphere xxx (2013) xxx–xxx

concentration of antibiotics present in the effluent from WWTP by approximately 94% in Australia (Watkinson et al., 2007), and Yang et al. reported low pressure membrane filtration, granular activated carbon adsorption, and ozonation can effectively remove the antibiotics in the primary effluent at high efficiencies in the USA (Yang et al., 2011). However, very limited information is available for wastewater reclamation plant (WRP) in China, where only few WWTPs use tertiary treatment or advanced treatment due to high cost and energy demand. In the present study, we investigated the occurrence and elimination of 22 antibiotics, including eight quinolones (QNs), nine sulfonamides (SAs) and five macrolides (MCs), from a wastewater reclamation plant (WRP) coupled with an ultrafiltration (UF) and ozonation system. The aim of the research was to estimate the antibiotic removal efficiencies during different treatment steps, to investigate the seasonal variation of selected antibiotics in the wastewater and sludge samples in the WRP, and to determine the risks of selected antibiotics to aquatic organisms. The results obtained in this study will have significant implications for the elimination of antibiotics from WWTPs and WRPs. 2. Materials and methods 2.1. Standards and reagents HPLC-grade methanol and acetonitrile were purchased from Fisher Scientific (Pittsburgh, PA, USA). Formic acid (98%) was purchased from Fluka. Ammonium formate (99%) and ammonium hydroxide (v/v, 50%) were purchased from Alfa Aesar. De-ionized (DI) water (>18.2 MX cm1) was prepared with the Milli-Q Advantage A10 system (Millipore, USA). Norfloxacin (NOR, 99.9%), ciprofloxacin (CIP, 99.9%), sarafloxacin (SAR, 95.0%), Ofloxacin (OFL, 99.9%), fleroxacin (FLE, 99.5%), lomefloxacin (LOM, 98.0%), difloxacin (DIF, 98.0%), enrofloxacin (ENR, 99.9%), sulfadiazine (SDZ, 99.7%), sulfamerazine (SMR, 99.9%), sulfadimethoxine (SDM, 99.4%), sulfisoxazole (SIA, 99.0%), sulfamonomethoxine (SMM, 99.0%), erythromycin (ERY, 99.1%), roxithromycin (ROX, 90.0%), josamycin (JOS, 98.0%), tylosin (TYL, 82.4%), and spiramycin (SPI, 88.9%) were purchased from Sigma– Aldrich (St. Louis, MO, USA). Sulfamethoxazole (SMX, 99.0%), sulfathiazole (STZ, 99.0%), sulfapyridine (SPD, 99.0%), and sulfamethazine (SMZ, 99.0%) were purchased from KaSei Industry Co., Ltd. (Tokyo, Japan). The following isotopically labelled compounds were used as surrogate standards at 100.0 lg L1 in methanol. Norfloxacin-d5 (NOR-d5), ofloxacin-d3 (OFL-d3) and sarafloxacin-d8 (SAR-d8) were purchased from Sigma–Aldrich (St. Louis, MO, USA). Sulfamethoxazole-d4 (SMX-d4), sulfamethazine-d4 (SMZ-d4), spiramycin I-d3 (SPI I-d3), and erythromycin-13C, d4 (ERY-13C, d4) were purchased from Toronto Research Chemicals (Oakville, ON, Canada). The physicochemical characteristics of the test antibiotics are listed in Table S1. 2.2. Sample collection The wastewater reclamation plant (Beijing, China) investigated in this work serves approximately 810 000 people with an average domestic wastewater flow of 200 000 m3 d1. This facility has a conventional activated sludge (CAS) system, coupled with a subsequent ultrafiltration and ozone oxidation system. Sewage is first treated with a screen, and then flows through the activated sludge system (anaerobic, anoxic, and oxic tanks). After a secondary clarification step, most of the effluent (120 000 m3 d1) is discharged into the receiving river, and the rest (80 000 m3 d1) is pumped into the ultrafiltration (UF) system and treated

in an ozonation tank. The UF system is equipped with six trains of ZeeWeed-1000 polyvinylidene fluoride (PVDF) membrane modules (pore size: 0.02 lm, Zenon Co., Canada). Each train contains nine cassettes of 60 modules per cassette. The module is operated in an outside-in configuration at a constant flow of 23 L (m2 h)1. For the ozonation process, ozone is produced continuously from three ozone generators (maximum ozone output: 8.4 kg h1. The ozone dosage and contact time in the reaction tank are 5 mg L1 and 15 min, respectively. Detailed information on this WRP is listed in Table 1. Seasonal samplings were carried out at the outlet of each treatment step in spring (April 15, 2011), summer (July 14, 2011), fall (November 17, 2011) and winter (January 8, 2011). The composite wastewater and sludge sampling of the WRP was taken as shown in Fig. 1. 24 h composite samples were collected in a flow proportional mode. At equal time increments (2 h), samples were collected and composited with volume proportionally to the flow rate by an automatic device at each sample point. All water samples were collected in 1-L amber glass bottles, which were washed with methanol and DI water before use. They were extracted as soon as possible after being filtered through 0.45 lm nylon membrane filters (Whatman, UK). Sludge samples were collected at the outlet of every treatment step in polyethylene bags. They were centrifuged and immediately freeze-dried after being delivered to the laboratory. The sludge was ground and sieved to smaller than 0.44 mm, and then stored at 20 °C until analysis. 2.3. Sample preparation and analysis Chemical analysis was performed following the EPA method 1694 with some modifications (EPA, 2007). Target antibiotics were extracted from water samples using AutoTrace SPE 280 (Dionex, USA) with an Oasis HLB cartridge (6 mL, 200 mg; Waters, USA), and from sludge samples with an ASE 350 pressurized liquid extraction (PLE) system (Dionex, Sunnyvale, CA, USA) and then purified using an HLB cartridge. Detailed information is listed in the Supplementary materials. 2.4. Liquid chromatography/tandem mass spectrometry The LC system was Dionex Liquid Chromatography Ultimate 3000 (Sunnyvale, CA, USA). An XTerra MS C18 column (3 lm, 100 mm  2 mm) was used as the analytical column at a flow rate of 0.20 mL min1. Methanol–acetonitrile (1:1, v/v) was used as mobile phase A, and 0.3% formic acid in water (containing 0.1% ammonium formate, v/v, pH = 2.9) was used as mobile phase B. The gradient program was as follows: the mobile phase starting conditions were 10% of A for 2.0 min, and A was increased to 70% in 10.0 min before being increased to 100% in 4.0 min; 100% of A for 3.0 min, followed by returning to the initial composition in 0.1 min, which was maintained for 13.9 min. The total run time was 33.0 min. The MS system consisted of a triple-quadrupole mass spectrometer (API 3200; Applied Biosystems/MDS SCIEX, US) with electrospray ionization (ESI). The instrument was operated in the positive electrospray ionization and multiple reactions monitoring (MRM) mode. The MS/MS parameters were optimized as follows: curtain gas pressure, 0.14 MPa; collision gas pressure, 0.02 MPa; ion spray voltage, 5000 V; temperature, 600 °C; gas 1, 0.38 MPa; and gas 2, 0.45 MPa. Other parameters of MS/MS and ion pair are listed in Table S2. 2.5. Quantification and quality control The method detection limits (MDLs) for antibiotics were determined as the minimum detectable concentration of an analyte

Please cite this article in press as: Li, W., et al. Occurrence and removal of antibiotics in a municipal wastewater reclamation plant in Beijing, China. Chemosphere (2013), http://dx.doi.org/10.1016/j.chemosphere.2013.01.040

3

W. Li et al. / Chemosphere xxx (2013) xxx–xxx Table 1 Treatment parameters in studied WRP.

a b c d e f

Parameters

Influent

Anaerobic

Anoxic

Oxic

Secondary effluent

Ultrafiltration effluent

Tertiary effluent

Return sludge

Excess sludge

Recirculated sludge

Water flow (105 m3 d1) Sludge flow (105 kg d1) pH HRTa (h) SRTb (d) TOCc (sludge, mg kg1) DOCd (water, mg L1) BOD5e (mg L1) CODcrf (mg L1) NH4–N (mg L1) TP (mg L1)

2 – 7.93 – –

4 17.8  105 6.69 1.5

10 44.7 6.67 3

10 43.4 7.59 10.8

0.8 – 7.35 – –

0.8 – 7.07 – –

216 77.8 – – – –

198 65.7 – – – –

208 49.1 – – – –

2 – 7.15 – – 185 25.7 8.99 41.1 2.98 0.744

– – 27.9 – –

– <6 <15 <1.5 <0.3

1.92 17.43 – – 20–25 217 – – – – –

0.08 0.5 – – – – – – – – –

6 25.6 – – – – – – – – –

– 210 427 49.3 6.29

HRT = hydraulic residence time. SRT = solid residence time. TOC = total organic carbon. DOC = dissolved organic carbon. BOD5 = five-day biochemical oxygen demand. CODcr = chemical oxygen demand consumption using the dichromate method.

Fig. 1. Schematic diagram of the treatment processes in the WRP and sampling site location.

with a signal-to-noise ratio of 3. The MDLs were 0.01–0.25 ng L1 in wastewater, and 0.02–0.5 lg kg1 in sludge samples. A calibration curve was generated across a wide range of concentrations (0.05–500 lg L1) and had strong linearity (r2 > 0.99). For each set of samples, at least one procedure blank and one independent check standard were run in sequence to check for background contamination and system performance. Correlation coefficients and limits of detection of the 22 antibiotics are listed in Table S3. The recoveries of SPE and PLE were shown in Table S4 in the Supplemental Information. For the spiking level of 20 lg L1, the recoveries values of SPE and PLE range from 84.7–108.1% and 71.1–122.5%, respectively. The results indicated that the SPE and PLE extraction is sufficient to support quantitative extraction. 2.6. Statistical analysis Statistical analyses were performed with IBM PASW Statistics 18.0 (SPSS Inc., 1993–2007). Pearson correlation analysis was used for correlation analysis. It was considered as statistically significant difference when p < 0.05 for this test. In order to assess the contribution of each treatment unit in the WRP to antibiotic removal, the relative fraction (%) of elimination contribution (EC) of antibiotics from the aqueous phase in each treatment unit in the WRP was calculated with the following equation:

EC ¼ ðC Inflow  C Outflow Þ=ðC Influent  C Tereff Þ  100%

ð1Þ

where CInflow and COutflow are the average concentrations of target compounds in inflow and outflow of each treatment unit, respectively; while CInfluent and CTereff are the concentrations of target compounds in influent and tertiary effluent of the WRP, respectively. Overall removal efficiencies of the target compounds in the conventional treatment (RC) and advance treatment (RA) were calculated as relative amounts to the influent concentrations using Eqs. (2) and (3), respectively:

RC ¼ ðC Influent  C Seceff Þ=C Influent  100%

ð2Þ

RA ¼ ðC Influent  C Tereff Þ=C Influent  100%

ð3Þ

where CInfluent, CSeceff and CTereff are the concentrations measured in influent, secondary effluent, and tertiary effluent of the WRP, respectively. Mass flux (W) was calculated using as the following equation:

W ¼ C dissolv ed  Q water þ C adsorbed  Q sludge

ð4Þ

where W is the total mass of individual antibiotic in aqueous and sorbed phase; Cdissolved and Cadsorbed represent the dissolved and adsorbed concentrations, respectively; Qwater and Qsludge represent the water and sludge daily flow rates during the sampling period, respectively. Fig. S1 shows the mass flux (g/d) of total antibiotics in WRP processes. To assess the mass variations of antibiotics under different treatment processes, the fraction (%) of mass loss (RLost) in each

Please cite this article in press as: Li, W., et al. Occurrence and removal of antibiotics in a municipal wastewater reclamation plant in Beijing, China. Chemosphere (2013), http://dx.doi.org/10.1016/j.chemosphere.2013.01.040

4

W. Li et al. / Chemosphere xxx (2013) xxx–xxx

treatment unit during activated sludge treatment was calculated using as the following equation:

RLost ¼ ðW Inflow  W Outflow Þ=W Inflow  100%

ð5Þ

where WInflow and WOutflow represent the mass flux of target antibiotic in inflow and outflow of each treatment unit. In general, losses below 20% are too low to suggest clearly degradation or biotransformation of a chemical (Katsoyiannis and Samara, 2005). In order to assess the contribution of sorption and degradation of the antibiotics in the WRP during activated sludge treatment, the mass load of antibiotics that was lost due to sum of all transformation processes WLost (g/d) was calculated using the following equation:

W Lost ¼ W Influent  W Effluent  W Sludge

ð6Þ

where WInfluent and WEffluen, respectively, represent the mass load of target antibiotic in influent and secondary effluent (g/d); WSludge is the mass output in excess sludge (g/d). 2.7. Risk characterization Hazard quotients (HQs) for the aquatic environment were calculated using the following formula:

HQ ¼ MEC=PNEC

ð7Þ

where MEC is the maximum measured environmental concentration, and PNEC is the predicted no effect concentration in water. PNEC was calculated following the formula:

PNEC ¼ ðLC50 or EC50 Þ=AF

ð8Þ

where LC50 or EC50 is the lowest median effective concentration value obtained from available literature, and AF is an appropriate standard assessment factor (1000) (Park and Choi, 2008; Escher et al., 2011). If the value of HQ > 1, the ecological impact is expected for the selected antibiotics (Brain et al., 2004; Hernando et al., 2006). 3. Results and discussion 3.1. Concentrations of antibiotics The concentrations of target antibiotics in the wastewater and sludge samples taken at various stages of the treatment are summarized in Tables 2 and 3, respectively. 3.1.1. Influent As shown in Table 2, a total of 17 antibiotics, including eight quinolones (NOR, CIP, DIF, ENR, FLE, OFL, LOM, and SAR), four sulfonamides (SMX, SPD, SMZ, and SDZ), and five macrolides (SPI, JOS, TYL, ROX and ERY), were detected in influent samples. The concentrations of other five sulfonamides STZ, SDM, SMR, SIA, and SMM were below MDLs in all samples. Their absence may be well explained by the fact that these drugs are primarily used for animals in Beijing. Among the three classes of antibiotics detected in influents, the total concentration of quinolones (mean 4916 ng L1) was higher than those of sulfonamides (mean 2961 ng L1) and macrolides (mean 365 ng L1). OFL was the dominant antibiotic in the influent, followed by SDZ, NOR, SMX, SPD, ERY, FLE and ROX. The concentrations of main target antibiotics detected in the influent in this study were comparable to those previously reported in Chongqing, China (OFL: 780 ng L1; NOR: 859 ng L1; SDZ: 1382 ng L1; ERY: 206 ng L1) (Chang et al., 2010), and USA (OFL: 400–1000 ng L1; SMX: 390–1000 ng L1) (Brown et al., 2006). These results indicated that the use patterns of these antibiotics were similar in these regions. However, the concentrations of

these antibiotics in the influent were at least ten times higher than those in Sweden (OFL: 9–30 ng L1; NOR: 18–27 ng L1) (Zorita et al., 2009), and Finland (OFL, mean: 100 ng L1; NOR, mean: 120 ng L1) (Vieno et al., 2007), indicating that the level of antibiotic pollution in Beijing was relative high. 3.1.2. Secondary effluent After the conventional treatment with activated sludge process (anaerobic–anoxic–oxic process), the 17 antibiotics were still detectable in secondary effluent samples. Among the three classes of antibiotics detected in secondary effluents, the total concentration of quinolones (mean 1869 ng L1) was higher than those of sulfonamides (mean 1053 ng L1) and macrolides (mean 353 ng L1). OFL was also the dominant antibiotic in the secondary effluent, followed by NOR, SDZ, SPD, SMX, ERY and ROX. These values were slightly higher than or similar to those reported in Hangzhou, China (OFL: 429 ng L1; NOR: 96 ng L1) (Tong et al., 2011), USA (OFL: 110 ng L1; SMX: 310 ng L1) (Brown et al., 2006), and South Korean (SMX, mean: 136 ng L1; ERY, mean: 130 ng L1) (Kim et al., 2007), but were much higher than those found in Finland (OFL, mean: 14 ng L1; NOR, mean: <24 ng L1) (Vieno et al., 2007). 3.1.3. Tertiary effluent After the advance treatment with ultrafiltration and ozone oxidation, only 14 antibiotics were detectable in tertiary effluent samples. Among the three classes of antibiotics detected in tertiary effluents, the total concentration of quinolones (mean 123 ng L1) was higher than those of sulfonamides (mean 25.9 ng L1) and macrolides (mean 24.7 ng L1). OFL was the dominant antibiotic in the tertiary effluent, followed by NOR, ERY, SMX, SDZ and ROX. In combination, the six antibiotics accounted for 92.1% of the total antibiotics in tertiary effluent. Limited information is available regarding the presence of antibiotics in the tertiary effluents or recycled water. In the present study, the concentrations of NOR and ERY were much higher than those reported in the tertiary effluents in Australia (NOR, median: 5 ng L1) (Watkinson et al., 2007) and USA (ERY, mean: 2 ng L1) (Yang et al., 2011), while the concentrations of SMX were much lower than those found in USA (SMX, mean: 80 ng L1) (Yang et al., 2011). 3.1.4. Sludge As shown in Table 3, 16 out of the 22 target antibiotics were detected in the sludge samples. The composition profiles of the detected antibiotics in different sludge samples were similar, because large amounts of sludge in the STP were cycled. This is consistent with those found in other treatment plants where activated sludge was used as secondary treatment process (Fan et al., 2011). Quinolones were the predominant antibiotics in the sludge, and the concentration of total quinolones (mean 9200 lg kg1) was two orders of magnitude higher than those of total sulfonamides (mean 63.7 lg kg1) and macrolides (mean 32.7 lg kg1). NOR and OFL were dominant and accounted for over 95% of the total antibiotics in the sludge samples. The concentrations of the two antibiotics detected in this study were similar to those found in Sweden (NOR: 4200 lg kg1) (Lindberg et al., 2005) and Italy (OFL: 3408 lg kg1) (Zuccato et al., 2010), while were much higher than those found in Japan (NOR: 87 lg kg1) (Okuda et al., 2009) and Spain (OFL: 8.4 lg kg1) (Radjenovic´ et al., 2009a). Sorption coefficient (Kd) is used to describe the reversible sorptive exchange of chemicals between water and sludge (Tolls, 2001). Kd is calculated according to the following equation: Kd = Cs/Cw, where Cw is the average concentration in water (ng L1), and Cs is the average concentration in sludge (lg kg1). The average Kd values of each antibiotic in anaerobic tank, anoxic tank, oxic tank, and

Please cite this article in press as: Li, W., et al. Occurrence and removal of antibiotics in a municipal wastewater reclamation plant in Beijing, China. Chemosphere (2013), http://dx.doi.org/10.1016/j.chemosphere.2013.01.040

5

W. Li et al. / Chemosphere xxx (2013) xxx–xxx Table 2 Summary of average concentrations of targeted antibiotics measured in composite water samples (ng L1). Analytes

NOR CIP DIF ENR FLE OFL LOM SAR STZ SMX SIA SPD SDM SMZ SDZ SMR SMM SPI JOS TYL ROX ERY QNs SAs MCs

Influent

Anaerobic

Anoxic

Oxic

Secondary clarifier

Ultrafiltration

Ozonation

Mean

Min

Max

Mean

Min

Max

Mean

Min

Max

Mean

Min

Max

Mean

Min

Max

Mean

Min

Max

Mean

Min

Max

1813 82.1 2.39 3.55 147 2794 63.7 9.37 n.d. 496 n.d. 451 n.d. 6.32 2009 n.d. n.d. 7.46 0.86 6.42 129 221 4916 2961 365

1368 35.0 n.d. n.d. n.d. 1445 40.4 n.d. n.d. 332 n.d. 281 n.d. 4.63 760 n.d. n.d. 3.08 n.d. 2.62 25.0 48.6 3703 1732 79.9

2746 119 9.55 7.92 375 3675 97.4 23.3 n.d. 646 n.d. 608 n.d. 7.95 4820 n.d. n.d. 11.0 2.11 17.2 224 520 6704 6080 758

1889 46.4 0.91 3.19 49.3 2746 75.4 8.71 n.d. 561 n.d. 569 n.d. 7.87 1579 n.d. n.d. 3.37 0.66 2.06 82.4 157 4818 2716 246

710 15.7 n.d. n.d. n.d. 796 27.5 n.d. n.d. 289 n.d. 409 n.d. 7.05 1080 n.d. n.d. n.d. n.d. 1.26 12.3 42 1701 2131 55.8

4266 82.0 3.06 11.7 104 4580 191 24.8 n.d. 956 n.d. 646 n.d. 9.60 2460 n.d. n.d. 8.82 2.08 3.52 218 394 8980 4069 622

1946 28.5 0.77 2.27 82.1 2261 59.4 9.50 n.d. 327 n.d. 495 n.d. 6.36 807 n.d. n.d. 3.06 0.13 2.11 100 172 4390 1635 278

552 18.2 n.d. n.d. n.d. 515 17.5 n.d. n.d. 113 n.d. 423 n.d. 3.89 303 n.d. n.d. 0.30 n.d. 1.52 17.1 27.4 1188 996 54.0

4481 44.6 3.06 3.26 213 4275 134 19.6 n.d. 482 n.d. 555 n.d. 8.39 1320 n.d. n.d. 5.75 0.53 2.83 302 490 9171 2273 800

1472 29.8 0.49 3.10 59.6 1971 53.5 5.22 n.d. 235 n.d. 330 n.d. 3.41 535 n.d. n.d. 3.08 0.59 1.99 66.5 129 3594 1104 201

317 5.85 n.d. n.d. n.d. 163 6.60 n.d. n.d. 34.9 n.d. 188 n.d. 1.78 162 n.d. n.d. n.d. n.d. 0.94 15.6 8.24 493 389 44.8

2976 54.3 1.97 7.20 125 3065 85.4 12.0 n.d. 313 n.d. 540 n.d. 4.18 904 n.d. n.d. 6.16 2.09 3.11 169 320 6265 1746 496

650 17.8 0.26 0.89 33.1 1140 21.0 5.27 n.d. 224 n.d. 243 n.d. 4.59 581 n.d. n.d. 3.27 0.28 1.80 169 178 1869 1053 353

234 2.08 n.d. n.d. n.d. 651 4.32 n.d. n.d. 133 n.d. 170 n.d. 2.48 234 n.d. n.d. 2.66 n.d. 0.90 37.6 44.6 898 627 94.2

1162 31.0 1.03 2.24 83.6 1561 35.2 13.0 n.d. 302 n.d. 306 n.d. 6.24 904 n.d. n.d. 4.82 1.08 2.80 530 522 2545 1518 1061

787 14.3 0.27 1.23 21.8 1129 29.7 6.16 n.d. 241 n.d. 218 n.d. 4.55 513 n.d. n.d. 3.57 0.23 1.70 143 186 1989 976 335

236 7.36 n.d. n.d. n.d. 593 10.0 n.d. n.d. 141 n.d. 127 n.d. 3.36 212 n.d. n.d. 1.81 n.d. 0.92 41.8 48.8 847 485 107

1564 23.2 1.06 2.94 52.0 1514 57.0 14.4 n.d. 290 n.d. 318 n.d. 5.98 874 n.d. n.d. 7.70 0.87 2.60 413 526 2994 1483 950

40.1 1.77 0.20 0.55 3.75 72.6 3.30 0.44 n.d. 12.3 n.d. 2.72 n.d. 0.10 10.7 n.d. n.d. n.d. n.d. n.d. 9.99 14.7 123 25.9 24.7

n.d. n.d. n.d. n.d. n.d. 19.4 n.d. n.d. n.d. 4.82 n.d. 0.94 n.d. n.d. 5.14 n.d. n.d. n.d. n.d. n.d. 0.43 n.d. 48.9 10.9 0.59

124 4.98 0.79 1.26 8.38 185 11.5 1.76 n.d. 20.2 n.d. 3.98 n.d. 0.39 19.0 n.d. n.d. n.d. n.d. n.d. 32.7 42.1 199 43.2 74.8

n.d.: not detected; Min: minimum; Max: maximum.

Table 3 Summary of average concentrations of targeted antibiotics measured in sludge samples (lg kg1, dw). Analytes

NOR CIP DIF ENR FLE OFL LOM SAR STZ SMX SIA SPD SDM SMZ SDZ SMR SMM SPI JOS TYL ROX ERY QNs SAs MCs

Anaerobic

Anoxic

Oxic

Return sludge

Average

Mean

Min

Max

Mean

Min

Max

Mean

Min

Max

Mean

Min

Max

Mean

Min

Max

6025 130 0.99 15.2 105 3820 130 2.99 0.00 9.47 0.07 19.3 0.00 0.30 47.2 0.00 0.00 0.00 0.00 0.00 30.2 9.77 10229 76.3 40.0

2570 44.4 0.00 5.18 57.9 1466 55.6 0.00 0.00 1.98 0.00 13.4 0.00 0.00 30.6 0.00 0.00 0.00 0.00 0.00 1.25 0.00 4204 47.9 1.25

9302 224 3.96 26.6 215 5740 238 7.90 0.00 25.3 0.28 27.7 0.00 0.54 67.4 0.00 0.00 0.00 0.00 0.00 72.1 25.4 15135 121 97.5

3887 97.7 1.11 10.8 77.9 2740 99.0 0.60 0.00 9.65 0.27 15.5 0.00 0.28 38.3 0.00 0.00 0.00 0.00 0.00 27.7 9.09 6914 63.9 36.8

1830 20.4 0.00 3.21 36.3 1006 31.7 0.00 0.00 3.24 0.00 10.6 0.00 0.00 22.3 0.00 0.00 0.00 0.00 0.00 4.11 0.00 2929 39.1 4.11

6260 235 3.69 18.2 147 6090 246 1.46 0.00 24.0 0.85 23.6 0.00 0.81 63.6 0.00 0.00 0.00 0.00 0.00 67.7 28.6 11031 112 96.3

5415 110 0.00 13.9 91.9 3695 117 2.87 0.00 10.5 0.00 16.6 0.00 0.17 41.7 0.00 0.00 0.60 0.00 0.00 29.4 7.75 9446 68.9 37.8

3887 47.3 0.00 7.59 55.1 2110 48.4 0.00 0.00 2.02 0.00 10.7 0.00 0.09 21.3 0.00 0.00 0.00 0.00 0.00 1.46 0.00 7001 42.1 3.85

7102 223 0.00 21.3 163 5760 220 9.31 0.00 26.0 0.00 25.5 0.00 0.26 66.9 0.00 0.00 2.39 0.00 0.00 80.6 24.8 11644 119 105

6386 101 3.93 16.4 129 3233 96.5 8.50 0.00 4.65 0.10 15.4 0.00 0.24 25.3 0.00 0.00 0.00 0.00 0.00 10.2 1.67 9975 45.8 11.9

2847 81.5 0.00 12.7 69.5 2056 56.9 0.00 0.00 1.03 0.00 10.2 0.00 0.12 17.1 0.00 0.00 0.00 0.00 0.00 2.55 0.00 5189 31.3 2.55

9852 121 11.80 21.0 211 5273 169 25.5 0.00 9.05 0.30 22.3 0.00 0.42 35.9 0.00 0.00 0.00 0.00 0.00 20.3 2.72 15667 59.3 22.6

5462 110 1.26 15.2 99.5 3400 109 3.21 0.00 9.13 0.10 16.4 0.00 0.24 37.8 0.00 0.00 0.30 0.00 0.00 24.9 7.24 9200 63.7 32.5

1830 20.4 0.00 3.21 36.3 1006 31.7 0.00 0.00 1.03 0.00 10.2 0.00 0.00 17.1 0.00 0.00 0.00 0.00 0.00 1.25 0.00 2929 31.3 1.25

9852 235 11.80 26.6 215 6090 246 25.5 0.00 26.0 0.85 27.7 0.00 0.81 67.4 0.00 0.00 2.39 0.00 0.00 80.6 28.6 15667 121 105

n.d.: not detected; Min: minimum; Max: maximum.

secondary clarifier are shown in Table 4. In present study, Kd of quinolones in sludge varied from 1051 to 8058 L kg1, suggesting that these antibiotics had tendencies to accumulate in the sludge. This is consistent with previous study that NOR, OFL, and CIP exhibited high adsorption potential onto activated sludge, with Kd values in the same range (2256.9–5122.7 L kg1) (Li and Zhang, 2010). Even higher Kd values of NOR (15 800 L kg1) and CIP (19 900 L kg1) were reported by (Golet et al., 2003), which found that sorption of quinolones in sludge was very strong. In contrast,

the four sulfonamides (SMX, SPD, SDZ, and SMZ) and the three macrolides (SPI, ERY, and ROX) had low Kd values ranging from 32.6 to 352 L kg1, and had little sorption affinities to sludge. The Kd values obtained in this study were comparable with those reported in several studies for SMX (32–77 L kg1), SMZ (13 L kg1), and ERY (74 L kg1) (Radjenovic´ et al., 2009b; Yu et al., 2011). Jones et al. (2002), Sipma et al., (2010) and Le-Minh et al. (2010) found the Kd of SMX (114–400 L kg1), SPD (200–398 L kg1), and ERY (160 L kg1) to be one order of magnitude higher compared to

Please cite this article in press as: Li, W., et al. Occurrence and removal of antibiotics in a municipal wastewater reclamation plant in Beijing, China. Chemosphere (2013), http://dx.doi.org/10.1016/j.chemosphere.2013.01.040

6

W. Li et al. / Chemosphere xxx (2013) xxx–xxx

Table 4 Solid–water partition coefficients (Kd, L kg1) of detected antibiotics in the activated sludge treatment units.

will be eliminated more efficiently in summer and fall (Castiglioni et al., 2006; Vieno et al., 2007).

Analytes

Anaerobic

Anoxic

Oxic

Secondary clarifier

Average

3.3. Removal of targeted antibiotics by various unit processes

NOR CIP DIF ENR FLE OFL LOM SAR STZ SMX SIA SPD SDM SMZ SDZ SMR SMM SPI JOS TYL ROX ERY QNs SAs MCs

3189 2798 1093 4777 2127 1391 1728 343  16.9  33.9  38.4 29.9  0.00 0.00 0.00 0.00 366 62.1 2123 28.1 164

1997 3423 1444 4757 949 1212 1666 62.7  29.5  31.2  44.4 47.4   0.00 0.00 0.00 276 52.8 1575 39.1 134

3680 3707 0.00 4494 1543 1875 2188 550  44.5  50.4  50.9 77.8   194 0.00 0.00 443 60.3 2628 62.4 189

9819 5652 11919 18437 3907 2836 4602 1613  20.7  63.5  52.5 43.6   0.00 0.00 0.00 60.3 9.35 5339 43.5 34.7

7693 8058 2675 7811 2119 2827 4575 1051  32.6  47.4  48.4 67.1   50.9 0.00 0.00 352 82.2 4586 49.1 197

Using the concentrations of selected antibiotics across the WRP (Table 2), we calculated the elimination contribution of the selected antibiotics from the aqueous phase in each step (i.e., anaerobic tank, anoxic tank, oxic tank, secondary clarifier, ultrafiltration, and ozonation) (Table 5). The overall removal efficiency of conventional treatment (RC) and advance treatment (RA) are presented in Table 6.

the present study, but also reported Kd for ROX (158–501 L kg1) in the same level as found in this study. These are in accordance with the conclusions from previous studies that some sulfonamides and macrolides are usually relatively mobile, and that macrolides have very low adsorption potential onto soil and sludge (Beausse, 2004; Göbel et al., 2005; Batt et al., 2007). Kds varied significantly among different tanks in the CAS system (Table 4). It has been reported that the adsorption of some antibiotics is related to organic carbon (Zhou et al., 2011; Jia et al., 2012). Therefore, the contents of total organic carbon (TOC, mg kg1) in sludge samples and dissolved organic carbon (DOC, mg L1) in water samples from the four treatment processes were investigated to assess the variation of Kds. The TOC values for sludge and the DOC values for wastewater samples collected from anaerobic, anoxic, oxic units and secondary clarifier (return sludge) are shown in Table 1. No significant correlation was found between TOC and Kd or between DOC and Kd of most detected antibiotics (Table S5). However, significant correlation was found between the ratio of TOC to DOC and Kd of six antibiotics detected in the four tanks. This result showed that the Kd of most antibiotics was positively correlated with TOC/DOC in the CAS system, indicating that organic carbon may influence the distribution of antibiotics between water and particulate matter to some extent. 3.2. Seasonal variation Seasonal variations of antibiotic concentrations were observed in water and sludge samples. As shown in Table S6, the concentrations of antibiotics in most samples were higher in spring and winter than those in summer and fall. It has been reported that the use of antibiotics in the winter may be higher than that in the summer (McArdell et al., 2003), so the seasonal variation of the concentrations of antibiotics may be due to more usages during the cold period of the year in this area. In addition, the temperature of the wastewater in summer and fall was much higher than that in spring and winter, which may facilitate the biodegradation of these antibiotics (Xu et al., 2011). Furthermore, due to higher microbial activities, antibiotics that are mainly removed by biodegradation

3.3.1. Conventional treatment 3.3.1.1. Quinolones. The overall removal efficiency of total quinolones by conventional treatment was 62.0% (Table 6). The mass balance results for detected antibiotic were expressed in mass fractions (%) in secondary effluent, excess sludge, and lost (biodegraded or transformed during activated sludge treatment) relative to the mass load in influent (Fig. 2). The calculated fractions of mass losses due to sorption and output of excess sludge for quinolones accounted for 44% of initial loadings, while the contribution of degradation was much less (<18%). In the activated sludge system, low mass changes were found in anaerobic (0%), anoxic (19%), oxic treatment units (26%), and secondary clarifier (6%) (Fig. S2), suggesting that biodegradation was of minor importance in the removal of quinolones. In addition, the data in Table 4 showed high Kd values (average: 642–8116 L kg1) of quinolones, indicating that sorption to sludge was a major pathway for the removal of these antibiotics from wastewater (Conkle et al., 2010). Several studies have also reported that the predominant removal mechanism of quinolones in the WWTPs is adsorption to sludge (Giger et al., 2003; Batt et al., 2007). For example, approximately 80% of the total mass of both NOR and CIP are absorbed to particles in the raw sewage water (Lindberg et al., 2006). The Kd value for each unit was 2123, 1575, 2628, and 5339 L kg1, respectively. As shown in Tables 4 and 5, the Kd value and elimination contribution of quinolones from the aqueous phase in secondary clarifier (Kd: 5339 L kg1; EC: 36.0%) were much higher than those in anaerobic (Kd: 2123 L kg1; EC: 2.03%), anoxic (Kd: 1575 L kg1; EC: 8.94%), and oxic tank (Kd: 2628 L kg1; EC: 16.6%), indicating that the adsorption of quinolones to sludge mainly occurred in this unit. It should be noted that the elimination contribution of quinolones in anaerobic tank was much lower than those in anoxic and oxic tanks. This result is consistent with the report that quinolones are stable in the anaerobic sludge, and no significant removal of the compounds occur in the anaerobic tank (Radjenovic´ et al., 2009b). Additionally, the elimination contribution might be influenced by contact time, so the lower values in anaerobic treatment could be explained by the hydraulic retention time (HRT) in this tank (1.5 h), which was shorter than those in anoxic tank (3.0 h) and oxic tank (10.8 h) (Table 1). The overall removal efficiency of individual quinolone by the conventional treatment is shown in Table 5. In our study, most quinolone antibiotics except for SAR (44%) were efficiently eliminated (59–86%) during the conventional treatment. These results were similar to those found in Italy (CIP: 60–63%; OFL: 43–57%) (Castiglioni et al., 2006) and South China (NOR: 70%; OFL: 57%) (Xu et al., 2007), while were lower than those reported in Finland (CIP: 84%; OFL: 83%) (Vieno et al., 2007), Sweden (CIP: 90%; OFL: 56%) (Zorita et al., 2009), and Switzerland (CIP: 79–86%; NOR: 80–87%) (Golet et al., 2002). 3.3.1.2. Sulfonamides. The overall removal efficiency of total sulfonamides by conventional treatment was 64% (Table 6). The calculated fractions of mass losses due to degradation for

Please cite this article in press as: Li, W., et al. Occurrence and removal of antibiotics in a municipal wastewater reclamation plant in Beijing, China. Chemosphere (2013), http://dx.doi.org/10.1016/j.chemosphere.2013.01.040

7

W. Li et al. / Chemosphere xxx (2013) xxx–xxx Table 5 Elimination contribution (%) for each antibiotic along the treatment units. Analytes

Conventional treatment

NOR CIP DIF ENR FLE OFL LOM SAR STZ SMX SIA SPD SDM SMZ SDZ SMR SMM SPI JOS TYL ROX ERY QNs SAs MCs

Advanced treatment

Anaerobic

Anoxic

Aerobic

Secondary clarifier

Ultrafiltration

Ozonation

4 44 68 12 68 2 19 7  14  26  25 22   55 23 68 39 31 2 8 35

3 22 6 31 23 18 26 9  48  17  24 39   4 62 1 15 7 9 37 9

27 2 13 28 16 11 10 48  19  37  47 14   0 53 2 29 21 17 18 23

46 15 7 74 18 31 54 1  2  19  19 2   3 36 3 87 24 36 2 45

8 4 3 11 8 0 15 10  3  6  1 3   4 6 2 23 4 3 3 5

42 16 3 23 13 39 44 64  47  48  72 25   48 27 27 112 83 39 32 91

Table 6 Overall removal efficiency (%) for each antibiotic by conventional and advanced treatment. Analytes

Conventional treatment

Advanced treatment

NOR CIP DIF ENR FLE OFL LOM SAR STZ SMX SIA SPD SDM SMZ SDZ SMR SMM SPI JOS TYL ROX ERY QNs SAs MCs

64 78 86 75 78 59 67 44 0 55 0 46 0 27 71 0 0 56 67 72 32 20 62 64 3

98 98 92 85 98 97 95 95 0 98 0 99 0 99 100 0 0 100 100 100 92 93 98 99 93

sulfonamides accounted for 63% of initial loadings, while the contribution of sorption and output of excess sludge was much less (<2%) (Fig. 2). The relative low Kd values (27.9–49.7 L kg1) also indicated that sorption to sludge had a limited contribution to the removal of sulfonamides. In the activated sludge system, mass change percentages were 9%, 1%, 27%, and 11% in anaerobic tank, anoxic tank, and oxic tank and secondary clarifier, respectively (Fig. S2). The high mass change in the oxic tank suggested that sulfonamides were mainly biodegraded in this unit. This result is consistent with reports that significant removal of sulfonamides

is achieved by biological treatment (Carballa et al., 2006; Yang et al., 2011). The removal efficiency of individual sulfonamide is shown in Table 5. The result showed that SDZ (71%) was efficiently eliminated, while the other three sulfonamides SMZ (27%), SPD (46%), and SMX (55%) were poorly eliminated. Similar or higher removal efficiencies in the STP were reported in Spain (SDZ: 78–91%; SMX: 54–71%; SMZ: 16%; SPD: 29–43%) (Garcia-Galan et al., 2011) and in Italy (SMX: 71%) (Castiglioni et al., 2006). 3.3.1.3. Macrolides. The overall removal efficiencies of total macrolides by conventional treatment were only 3% (Table 6), which was consistent with previous results that conventional STPs had low removal efficiencies for macrolides (McArdell et al., 2003; Joss et al., 2006). The fractions of mass losses due to sorption and degradation for macrolides accounted for 4% and 1% of initial loadings, respectively (Fig. 2), showing that macrolides are stable in sewage during the treatment process. This is consistent with the results that the most of them are resistant to the processes carried out in STPs (Xu et al., 2007; Baquero et al., 2008). In the activated sludge system, mass change percentages were 1%, 17%, 18% and 10% in anaerobic, anoxic, aerobic treatment units, and secondary clarifier, respectively (Fig. S2). The low mass change also suggested that biodegradation in activated sludge treatment units was of minor importance in the removal of macrolides. Additionally, the Kd value for each unit was 164, 134, 189, and 34.7 L kg1, respectively. The results showed that macrolides had low adsorption potential, and sorption to sludge accounted for only minor contribution to the removal of most macrolides in STPs (Göbel et al., 2005). The removal efficiency of individual macrolide ranged from 32% (ROX) to 72% (TYL). The three 16-membered macrolides, SPI, JOS and TYL, which were detected with low frequencies and at relatively low concentrations, were removed effectively (56–72%) during the conventional treatment. However, the other two 14-membered macrolides, ROX and ERY, were poorly removed, with a removal efficiency of 32% and 20%, respectively. For ERY, similar removal efficiencies were obtained in STPs in Spain (35.4%) (Radjenovic´ et al., 2009b) and South China (26%) (Xu et al., 2007). It should be noted that the

Please cite this article in press as: Li, W., et al. Occurrence and removal of antibiotics in a municipal wastewater reclamation plant in Beijing, China. Chemosphere (2013), http://dx.doi.org/10.1016/j.chemosphere.2013.01.040

8

W. Li et al. / Chemosphere xxx (2013) xxx–xxx

Fig. 2. Mass fractions (%) of QNs, SAs and MCs detected in secondary effluent (WEffluent), excess sludge (WSludge), and lost (WLost) during activated sludge treatment.

concentration of ROX in secondary effluent was higher than that in influent. It has been suggested that negative removals of macrolides are likely due to the presence of deconjugation of conjugated metabolites during the treatment process (Xu et al., 2007). 3.3.2. Advance treatment processes The incomplete removal of antibiotics in the secondary effluent indicated relatively high release rates of selected antibiotics into the environment. In order to minimize environmental and human exposure, advance treatment processes may be necessary to further remove these compounds. 3.3.2.1. Ultrafiltration. As shown in Table 5, the antibiotic concentrations in effluent of ultrafiltration process were similar to or even higher than those in secondary effluents. The UF membranes had little elimination contributions (<10%) to the target antibiotics except for ROX (23%). This result is comparable to the results of previous studies that this process contributes less than 10% of the overall removal efficiency (Al-Rifai et al., 2011). Since UF membrane pores (0.04 mm) are significantly larger than the contaminant molecules, it is difficult to retain antibiotics by size exclusions. The little contribution might be due to the adsorption of these antibiotics on the membrane, or the removal by size exclusion of large particles on which antibiotics could be attached. Generally, the contribution of UF to antibiotic removal was negligible. It has been reported that the UF membrane retains typical hydrophobic antibiotics mainly through hydrophobic adsorption (Yoon et al., 2006), and the retention is in accordance with Log Kow value (Yoon et al., 2007). Considering the relative high Log Kow for ROX (2.1–2.8) among the target antibiotics (Table S1), its relative high removal efficiency (23%) in this process was reasonable. 3.3.2.2. Ozonation. It was obvious that ozonation effectively removed most of the remaining antibiotics in the ultrafiltration process (Table 5), with the excellent removal efficiencies ranging from 85% to 100%. This good performance of ozonation in the present study is in agreement with previous results that ozonation is responsible for most of the removal of antibiotics (Nakada et al., 2008). It has been shown that quinolones, sulfonamides, and macrolides are predominantly transformed via direct reaction with ozone during this process (Dodd et al., 2009; Reungoat et al., 2012). Fig. S1 lists three classes of antibiotics with the ozone-reactive moieties. Generally, many of them contain amino groups, which

are susceptible to chemical attack by ozone (Yang et al., 2011). The reactive tertiary amino groups, aniline moiety groups and tertiary dimethylamino groups are characteristics for quinolones, sulfonamides and macrolides, respectively (Lange et al., 2006; Nakada et al., 2007). The three 16-membered macrolides SPI, JOS and TYL were completely removed from the tertiary effluent, and their removal efficiencies were higher than those of the two 14-membered ROX (92%) and ERY (93%) (Table 5). These results may be explained by the fact that the 16-membered macrolides possess a diene moiety assigned to the reaction of ozone (Fig. S3). It should be noted that ENR showed the lowest removal (85%) in this process, probably due to the strong sorption to sludge particles, which might offer some protection against O3 attack and oxidation (Huber et al., 2005). Generally, the removals of most antibiotics by UF are poor, while ozonation appears to be more effective. Although the overall removal efficiencies for all antibiotics were excellent, this WRP was not enough to completely remove these antibiotics from secondary effluent. 3.4. Environmental risk assessment To study the occurrence of antibiotics in environmental water due to discharge from STPs, we investigated, on April 15, July 14 and November 17, 2011, the occurrence of these antibiotics in the receiving river, which was one of the main wastewater channels in Beijing. The sampling sites for the receiving river were located at 4- and 2-km upstream, and 2- and 4-km downstream of the secondary effluent outfall of WRP. As shown in Table S7, 15 antibiotics were detected in river samples, and the concentrations of these antibiotics in upstream and downstream varied slightly. It should be noted that DIF and SAR, which were occasionally detected in STP effluents at relatively low levels, were not observed in the receiving river. Due to the more effective removal of these compounds during conventional treatment, the concentrations of NOR, OFL, SDZ and SMX in the secondary effluent samples were much lower than those in the river samples. However, the concentrations of the poorly removed antibiotics in the secondary effluent samples, such as ROX and ERY, were much higher than those in the river samples. This implied that the inputs of antibiotics via secondary effluent had a great influence on the receiving waters. In the present study, some antibiotics were frequently found in secondary effluent and tertiary effluent. Several studies have re-

Please cite this article in press as: Li, W., et al. Occurrence and removal of antibiotics in a municipal wastewater reclamation plant in Beijing, China. Chemosphere (2013), http://dx.doi.org/10.1016/j.chemosphere.2013.01.040

W. Li et al. / Chemosphere xxx (2013) xxx–xxx

ported that antibiotics in the aquatic environment may pose potential risks to the aquatic ecosystems (Lindberg et al., 2007). Therefore, it is necessary to estimate the ecological risk of these antibiotics on organisms in these water samples. Hazard quotients (HQs) are usually calculated for environmental risk assessment. If the value of HQ is above 1, an ecological impact is expected for the selected antibiotics. A summary of the acute median effective concentrations (LC50 or EC50) of the nine antibiotics to different aquatic organisms (algae, plant, invertebrate and fish) is presented in Table S8. Based on the acute toxicity data, the HQs calculated using formula (4) are summarized in Table S9. After conventional treatment, high HQ values of OFL (74.3), ERY (26.1), SMX (10.1), SDZ (6.70) and CIP (1.82) were found in algae, and high HQ values of OFL (12.4), SMX (3.73) and NOR (1.27) were also found in plants in the secondary effluent, suggesting potential adverse ecological consequences on aquatic algae and plants. If we considered the secondary effluents for estimating concentrations in receiving waters with a dilution factor of 10 (Duong et al., 2008), the HQ values of OFL, ERY, and SMX were also more than 1, indicating that adverse effects on algae and plants were present in secondary wastewater effluent. After ultrafiltration and ozone oxidation, all antibiotics, except for OFL (algae: 8.79; plant: 1.47) and ERY (algae: 2.10), had HQ values <1 in tertiary effluent. The result showed that the risks of antibiotics significantly decreased after advance treatment. However, the risks of OFL and ERY on organisms required further investigations. 4. Conclusions In the present study, the occurrence and elimination of 22 antibiotics were investigated at a wastewater recycling plant in Beijing. Most antibiotics were detected in influent, secondary effluent, tertiary effluent and sludge samples. This study showed that the removal efficiencies of most target antibiotics by conventional treatment and UF were poor, while they could be effectively removed by ozonation oxidation. Risk assessments showed that risks of antibiotic contamination were significantly decreased after advance treatment. Although wastewater reclamation is a great approach to reduce the release of antibiotics into aquatic environment, the threat of these antibiotics in wastewater reuse should be of concern. Acknowledgements This work was supported by the National Natural Science Foundation of China (Nos. 20837003 and 20890111) and the National Basic Research Program of China (2009CB421605). Appendix A. Supplementary material Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.chemosphere. 2013.01.040. References Al-Rifai, J.H., Khabbaz, H., Schafer, A.I., 2011. Removal of pharmaceuticals and endocrine disrupting compounds in a water recycling process using reverse osmosis systems. Sep. Purif. Technol. 77, 60–67. Baquero, F., Martinez, J.L., Canton, R., 2008. Antibiotics and antibiotic resistance in water environments. Curr. Opin. Biotechnol. 19, 260–265. Batt, A.L., Kim, S., Aga, D.S., 2007. Comparison of the occurrence of antibiotics in four full-scale wastewater treatment plants with varying designs and operations. Chemosphere 68, 428–435.

9

Beausse, J., 2004. Selected drugs in solid matrices: a review of environmental determination, occurrence and properties of principal substances. Trac – Trend Anal. Chem. 23, 753–761. Brain, R.A., Johnson, D.J., Richards, S.M., Hanson, M.L., Sanderson, H., Lam, M.W., Young, C., Mabury, S.A., Sibley, P.K., Solomon, K.R., 2004. Microcosm evaluation of the effects of an eight pharmaceutical mixture to the aquatic macrophytes Lemna gibba and Myriophyllum sibiricum. Aquat. Toxicol. 70, 23–40. Brown, K., Kulis, J., Thomson, B., Chapman, T., Mawhinney, D., 2006. Occurrence of antibiotics in hospital, residential, and dairy effluent, municipal wastewater, and the Rio Grande in New Mexico. Sci. Total Environ. 366, 772–783. Carballa, M., Omil, F., Alder, A.C., Lema, J.M., 2006. Comparison between the conventional anaerobic digestion of sewage sludge and its combination with a chemical or thermal pre-treatment concerning the removal of pharmaceuticals and personal care products. Water Sci. Technol. 53, 109–117. Castiglioni, S., Bagnati, R., Fanelli, R., Pomati, F., Calamari, D., Zuccato, E., 2006. Removal of pharmaceuticals in sewage treatment plants in Italy. Environ. Sci. Technol. 40, 357–363. Chang, X., Meyer, M.T., Liu, X., Zhao, Q., Chen, H., Chen, J.-a., Qiu, Z., Yang, L., Cao, J., Shu, W., 2010. Determination of antibiotics in sewage from hospitals, nursery and slaughter house, wastewater treatment plant and source water in Chongqing region of three Gorge reservoir in China. Environ. Pollut. 158, 1444–1450. Clara, M., Strenn, B., Gans, O., Martinez, E., Kreuzinger, N., Kroiss, H., 2005. Removal of selected pharmaceuticals, fragrances and endocrine disrupting compounds in a membrane bioreactor and conventional wastewater treatment plants. Water Res. 39, 4797–4807. Conkle, J.L., Lattao, C., White, J.R., Cook, R.L., 2010. Competitive sorption and desorption behavior for three fluoroquinolone antibiotics in a wastewater treatment wetland soil. Chemosphere 80, 1353–1359. Dirany, A., Aaron, S.E., Oturan, N., Sires, I., Oturan, M.A., Aaron, J.J., 2011. Study of the toxicity of sulfamethoxazole and its degradation products in water by a bioluminescence method during application of the electro-Fenton treatment. Anal. Bioanal. Chem. 400, 353–360. Dodd, M.C., Kohler, H.P.E., Von Gunten, U., 2009. Oxidation of antibacterial compounds by ozone and hydroxyl radical: elimination of biological activity during aqueous ozonation processes. Environ. Sci. Technol. 43, 2498–2504. Duong, H.A., Pham, N.H., Nguyen, H.T., Hoang, T.T., Pham, H.V., Pham, V.C., Berg, M., Giger, W., Alder, A.C., 2008. Occurrence, fate and antibiotic resistance of fluoroquinolone antibacterials in hospital wastewaters in Hanoi, Vietnam. Chemosphere 72, 968–973. EPA, 2007. Method 1694: Pharmaceuticals and Personal Care Products in Water, Soil, Sediment, and Biosolids by HPLC/MS/MS. . Escher, B.I., Baumgartner, R., Koller, M., Treyer, K., Lienert, J., McArdell, C.S., 2011. Environmental toxicology and risk assessment of pharmaceuticals from hospital wastewater. Water Res. 45, 75–92. Fan, Z., Wu, S., Chang, H., Hu, J., 2011. Behaviors of glucocorticoids, androgens and progestogens in a municipal sewage treatment plant: comparison to estrogens. Environ. Sci. Technol. 45, 2725–2733. Gao, L., Shi, Y., Li, W., Niu, H., Liu, J., Cai, Y., 2012. Occurrence of antibiotics in eight sewage treatment plants in Beijing, China. Chemosphere 86, 665–671. Garcia-Galan, M.J., Diaz-Cruz, M.S., Barcelo, D., 2011. Occurrence of sulfonamide residues along the Ebro river basin: removal in wastewater treatment plants and environmental impact assessment. Environ. Int. 37, 462–473. Giger, W., Alder, A.C., Golet, E.M., Kohler, H.P.E., McArdell, C.S., Molnar, E., Siegrist, H., Suter, M.J.F., 2003. Occurrence and fate of antibiotics as trace contaminants in wastewaters, sewage sludges, and surface waters. Chimia 57, 485–491. Göbel, A., Thomsen, A., McArdell, C.S., Joss, A., Giger, W., 2005. Occurrence and sorption behavior of sulfonamides, macrolides, and trimethoprim in activated sludge treatment. Environ. Sci. Technol. 39, 3981–3989. Golet, E.M., Alder, A.C., Giger, W., 2002. Environmental exposure and risk assessment of fluoroquinolone antibacterial agents in wastewater and river water of the Glatt valley watershed, Switzerland. Environ. Sci. Technol. 36, 3645–3651. Golet, E.M., Xifra, I., Siegrist, H., Alder, A.C., Giger, W., 2003. Environmental exposure assessment of fluoroquinolone antibacterial agents from sewage to soil. Environ. Sci. Technol. 37, 3243–3249. Hernando, M., Mezcua, M., Fernandezalba, A., Barcelo, D., 2006. Environmental risk assessment of pharmaceutical residues in wastewater effluents, surface waters and sediments. Talanta 69, 334–342. Huber, M.M., GÖbel, A., Joss, A., Hermann, N., LÖffler, D., McArdell, C.S., Ried, A., Siegrist, H., Ternes, T.A., von Gunten, U., 2005. Oxidation of pharmaceuticals during ozonation of municipal wastewater effluents: a pilot study. Environ. Sci. Technol. 39, 4290–4299. Jia, A., Wan, Y., Xiao, Y., Hu, J., 2012. Occurrence and fate of quinolone and fluoroquinolone antibiotics in a municipal sewage treatment plant. Water Res. 46, 387–394. Jones, O.A.H., Voulvoulis, N., Lester, J.N., 2002. Aquatic environmental assessment of the top 25 English prescription pharmaceuticals. Water Res. 36, 5013–5022. Joss, A., Zabczynski, S., Göbel, A., Hoffmann, B., Löffler, D., McArdell, C.S., Ternes, T.A., Thomsen, A., Siegrist, H., 2006. Biological degradation of pharmaceuticals in municipal wastewater treatment: proposing a classification scheme. Water Res. 40, 1686–1696. Katsoyiannis, A., Samara, C., 2005. Persistent organic pollutants (POPs) in the conventional activated sludge treatment process: fate and mass balance. Environ. Res. 97, 245–257.

Please cite this article in press as: Li, W., et al. Occurrence and removal of antibiotics in a municipal wastewater reclamation plant in Beijing, China. Chemosphere (2013), http://dx.doi.org/10.1016/j.chemosphere.2013.01.040

10

W. Li et al. / Chemosphere xxx (2013) xxx–xxx

Kim, S.D., Cho, J., Kim, I.S., Vanderford, B.J., Snyder, S.A., 2007. Occurrence and removal of pharmaceuticals and endocrine disruptors in South Korean surface, drinking, and waste waters. Water Res. 41, 1013–1021. Lange, F., Cornelissen, S., Kubac, D., Sein, M.M., von Sonntag, J., Hannich, C.B., Golloch, A., Heipieper, H.J., Moder, M., von Sonntag, C., 2006. Degradation of macrolide antibiotics by ozone: a mechanistic case study with clarithromycin. Chemosphere 65, 17–23. Le-Minh, N., Khan, S.J., Drewes, J.E., Stuetz, R.M., 2010. Fate of antibiotics during municipal water recycling treatment processes. Water Res. 44, 4295–4323. Leung, H.W., Minh, T.B., Murphy, M.B., Lam, J.C.W., So, M.K., Martin, M., Lam, P.K.S., Richardson, B.J., 2012. Distribution, fate and risk assessment of antibiotics in sewage treatment plants in Hong Kong, South China. Environ. Int. 42, 1–9. Li, B., Zhang, T., 2010. Biodegradation and adsorption of antibiotics in the activated sludge process. Environ. Sci. Technol. 44, 3468–3473. Lin, A.Y.-C., Yu, T.-H., Lateef, S.K., 2009. Removal of pharmaceuticals in secondary wastewater treatment processes in Taiwan. J. Hazard. Mater. 167, 1163–1169. Lindberg, R.H., Wennberg, P., Johansson, M.I., Tysklind, M., Andersson, B.A.V., 2005. Screening of human antibiotic substances and determination of weekly mass flows in five sewage treatment plants in Sweden. Environ. Sci. Technol. 39, 3421–3429. Lindberg, R.H., Olofsson, U., Rendahl, P., Johansson, M.I., Tysklind, M., Andersson, B.A.V., 2006. Behavior of fluoroquinolones and trimethoprim during mechanical, chemical, and active sludge treatment of sewage water and digestion of sludge. Environ. Sci. Technol. 40, 1042–1048. Lindberg, R.H., Bjorklund, K., Rendahl, P., Johansson, M.I., Tysklind, M., Andersson, B.A.V., 2007. Environmental risk assessment of antibiotics in the Swedish environment with emphasis on sewage treatment plants. Water Res. 41, 613– 619. McArdell, C.S., Molnar, E., Suter, M.J.F., Giger, W., 2003. Occurrence and fate of macrolide antibiotics in wastewater treatment plants and in the Glatt valley watershed, Switzerland. Environ. Sci. Technol. 37, 5479–5486. Nakada, N., Shinohara, H., Murata, A., Kiri, K., Managaki, S., Sato, N., Takada, H., 2007. Removal of selected pharmaceuticals and personal care products (PPCPs) and endocrine-disrupting chemicals (EDCs) during sand filtration and ozonation at a municipal sewage treatment plant. Water Res. 41, 4373–4382. Nakada, N., Kiri, K., Shinohara, H., Harada, A., Kuroda, K., Takizawa, S., Takada, H., 2008. Evaluation of pharmaceuticals and personal care products as watersoluble molecular markers of sewage. Environ. Sci. Technol. 42, 6347–6353. Okuda, T., Yamashita, N., Tanaka, H., Matsukawa, H., Tanabe, K., 2009. Development of extraction method of pharmaceuticals and their occurrences found in Japanese wastewater treatment plants. Environ. Int. 35, 815–820. Onesios, K.M., Yu, J.T., Bouwer, E.J., 2009. Biodegradation and removal of pharmaceuticals and personal care products in treatment systems: a review. Biodegradation 20, 441–466. Park, S., Choi, K., 2008. Hazard assessment of commonly used agricultural antibiotics on aquatic ecosystems. Ecotoxicology 17, 526–538. Radjenovic´, J., Jelic´, A., Petrovic´, M., Barceló, D., 2009a. Determination of pharmaceuticals in sewage sludge by pressurized liquid extraction (PLE) coupled to liquid chromatography–tandem mass spectrometry (LC–MS/MS). Anal. Bioanal. Chem. 393, 1685–1695. Radjenovic´, J., Petrovic, M., Barcelo, D., 2009b. Fate and distribution of pharmaceuticals in wastewater and sewage sludge of the conventional

activated sludge (CAS) and advanced membrane bioreactor (MBR) treatment. Water Res. 43, 831–841. Reungoat, J., Escher, B.I., Macova, M., Argaud, F.X., Gernjak, W., Keller, J., 2012. Ozonation and biological activated carbon filtration of wastewater treatment plant effluents. Water Res. 46, 863–872. Samanidou, V.F., Evaggelopoulou, E.N., 2007. Analytical strategies to determine antibiotic residues in fish. J. Sep. Sci. 30, 2549–2569. Sebastine, I.M., Wakeman, R.J., 2003. Consumption and environmental hazards of pharmaceutical substances in the UK. Process. Saf. Environ. 81, 229–235. Sipma, J., Osuna, B., Collado, N., Monclus, H., Ferrero, G., Comas, J., Rodriguez-Roda, I., 2010. Comparison of removal of pharmaceuticals in MBR and activated sludge systems. Desalination 250, 653–659. Tamtam, F., van Oort, F., Le Bot, B., Dinh, T., Mompelat, S., Chevreuil, M., Lamy, I., Thiry, M., 2011. Assessing the fate of antibiotic contaminants in metal contaminated soils four years after cessation of long-term waste water irrigation. Sci. Total Environ. 409, 540–547. Tong, C., Zhuo, X., Guo, Y., 2011. Occurrence and risk assessment of four typical fluoroquinolone antibiotics in raw and treated sewage and in receiving waters in Hangzhou China. J. Agr. Food Chem. 59, 7303–7309. Vieno, N., Tuhkanen, T., Kronberg, L., 2007. Elimination of pharmaceuticals in sewage treatment plants in Finland. Water Res. 41, 1001–1012. Watkinson, A.J., Murby, E.J., Costanzo, S.D., 2007. Removal of antibiotics in conventional and advanced wastewater treatment: implications for environmental discharge and wastewater recycling. Water Res. 41, 4164–4176. Xu, W., Zhang, G., Li, X., Zou, S., Li, P., Hu, Z., Li, J., 2007. Occurrence and elimination of antibiotics at four sewage treatment plants in the Pearl River Delta (PRD), South China. Water Res. 41, 4526–4534. Xu, B., Mao, D., Luo, Y., Xu, L., 2011. Sulfamethoxazole biodegradation and biotransformation in the water–sediment system of a natural river. Bioresour. Technol. 102, 7069–7076. Yang, X., Flowers, R.C., Weinberg, H.S., Singer, P.C., 2011. Occurrence and removal of pharmaceuticals and personal care products (PPCPs) in an advanced wastewater reclamation plant. Water Res. 45, 5218–5228. Yoon, Y., Westerhoff, P., Snyder, S., Wert, E., 2006. Nanofiltration and ultrafiltration of endocrine disrupting compounds, pharmaceuticals and personal care products. J. Membr. Sci. 270, 88–100. Yoon, Y., Westerhoff, P., Snyder, S.A., Wert, E.C., Yoon, J., 2007. Removal of endocrine disrupting compounds and pharmaceuticals by nanofiltration and ultrafiltration membranes. Desalination 202, 16–23. Yu, T.H., Lin, A.Y.C., Panchangam, S.C., Hong, P.K.A., Yang, P.Y., Lin, C.F., 2011. Biodegradation and bio-sorption of antibiotics and non-steroidal antiinflammatory drugs using immobilized cell process. Chemosphere 84, 1216– 1222. Zhou, L., Ying, G., Zhao, J., Yang, J., Wang, L., Yang, B., Liu, S., 2011. Trends in the occurrence of human and veterinary antibiotics in the sediments of the Yellow River, Hai River and Liao River in northern China. Environ. Pollut. 159, 1877–1885. Zorita, S., Mårtensson, L., Mathiasson, L., 2009. Occurrence and removal of pharmaceuticals in a municipal sewage treatment system in the south of Sweden. Sci. Total Environ. 407, 2760–2770. Zuccato, E., Castiglioni, S., Bagnati, R., Melis, M., Fanelli, R., 2010. Source, occurrence and fate of antibiotics in the Italian aquatic environment. J. Hazard. Mater. 179, 1042–1048.

Please cite this article in press as: Li, W., et al. Occurrence and removal of antibiotics in a municipal wastewater reclamation plant in Beijing, China. Chemosphere (2013), http://dx.doi.org/10.1016/j.chemosphere.2013.01.040