Occurrence of pharmaceuticals in a municipal wastewater treatment plant: Mass balance and removal processes

Chemosphere 88 (2012) 17–24

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Occurrence of pharmaceuticals in a municipal wastewater treatment plant: Mass balance and removal processes Pin Gao a,b,⇑, Yunjie Ding c, Hui Li c, Irene Xagoraraki b a

College of Environmental Science and Engineering, Donghua University, Shanghai 201620, China Department of Civil and Environmental Engineering, Michigan State University, East Lansing, MI 48824, USA c Department of Crop and Soil Sciences, Michigan State University, East Lansing, MI 48824, USA b

a r t i c l e

i n f o

Article history: Received 2 November 2011 Received in revised form 1 February 2012 Accepted 2 February 2012 Available online 9 April 2012 Keywords: Pharmaceuticals Wastewater treatment Mass balance Biodegradation Sorption

a b s t r a c t Occurrence and removal efficiencies of fifteen pharmaceuticals were investigated in a conventional municipal wastewater treatment plant in Michigan. Concentrations of these pharmaceuticals were determined in both wastewater and sludge phases by a high-performance liquid chromatograph coupled to a tandem mass spectrometer. Detailed mass balance analysis was conducted during the whole treatment process to evaluate the contributing processes for pharmaceutical removal. Among the pharmaceuticals studied, demeclocycline, sulfamerazine, erythromycin and tylosin were not detected in the wastewater treatment plant influent. Other target pharmaceuticals detected in wastewater were also found in the corresponding sludge phase. The removal efficiencies of chlortetracycline, tetracycline, sulfamerazine, acetaminophen and caffeine were >99%, while doxycycline, oxytetracycline, sulfadiazine and lincomycin exhibited relatively lower removal efficiencies (e.g., <50%). For sulfamethoxazole, the removal efficiency was approximately 90%. Carbamazepine manifested a net increase of mass, i.e. 41% more than the input from the influent. Based on the mass balance analysis, biotransformation is believed to be the predominant process responsible for the removal of pharmaceuticals (22% to 99%), whereas contribution of sorption to sludge was relatively insignificant (7%) for the investigated pharmaceuticals. Ó 2012 Elsevier Ltd. All rights reserved.

1. Introduction Release of pharmaceuticals into the environment has received a lot of attention in recent years. A significant number of these tracelevel compounds have been frequently detected both in aqueous (wastewater, drinking water, surface water, and groundwater) and solid (sludge, soil, and sediments) samples (Christian et al., 2003; Göbel et al., 2005; Kim and Carlson, 2007; Kümmerer, 2009). Increasing evidence indicates possible adverse impacts to the target organisms due to long-term and low-dosed exposures to pharmaceuticals in the environment, including chronic toxicity, endocrine disruption, antibiotic resistance, as well as toxic effects on reproduction of terrestrial and aquatic organisms (Andreozzi et al., 2004; Fent et al., 2006; Allen et al., 2010). Antibiotics are a particularly important class of pharmaceuticals used extensively in human and veterinary medicine, for the purpose of preventing or treating diseases, increasing feed efficiency and improving growth rate for livestock (Sarmah et al., 2006). Due to the emergence of possible spread and maintenance of resistance in bacteria and pathogens, antibiotic occurrence and fate in ⇑ Corresponding author. Address: 2999 North Renmin Rd., Songjiang District, Shanghai 201620, China. Tel.: +86 21 67792558; fax: +86 21 67792522. E-mail address: [email protected] (P. Gao). 0045-6535/$ - see front matter Ó 2012 Elsevier Ltd. All rights reserved. doi:10.1016/j.chemosphere.2012.02.017

the natural environment is of great importance (Allen et al., 2010). After consumption by humans and animals, antibiotics can be partially metabolized, and the metabolites are then excreted to urine or feces. Kümmerer and Henninger (2003) reported that the metabolization rate of all antibiotics used in Germany was estimated to be 30%. In other words, approximately 70% of the total amount is excreted unchanged into raw wastewater. Since the WWTPs are not designed to effectively remove these micropollutants, the elimination of pharmaceuticals in conventional activated sludge treatment process is incomplete. As a result, pharmaceutical residues could enter the environment via treated effluent discharge and land application of biosolids. To date, most previous studies focused on the levels of pharmaceutical residues in the final effluent from WWTPs. The corresponding removal efficiency varied significantly for each pharmaceutical, for instance, from 17% (Rosal et al., 2010) to 98% (Peng et al., 2006) for sulfamethoxazole (SMX), from 12% (Spongberg and Witter, 2008) to 80% (Karthikeyan and Meyer, 2006) for tetracycline (TC), from 4.3% to 72% (Rosal et al., 2010) for erythromycin-H2O (ERY), and less than 30% for carbamazepine (CBZ) (Miao et al., 2005) and 17% for lincomycin (LCM) (Karthikeyan and Meyer, 2006). Meanwhile, pharmaceutical residues were frequently detected in biosolids at lg kg1 to mg kg1 levels. An average concentration of 68 lg kg1 dry weight (dw) for SMX in activated sludge was

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reported by Göbel et al. (2005), and about 15 lg kg1 dw for TC in biosolids was shown by Spongberg and Witter (2008). Moreover, SMX concentrations as high as 1020 lg L1 and 220 lg L1 were reported to present in surface water and groundwater (Lindsey et al., 2001). Continuous input of low-level pharmaceuticals from WWTPs can adversely affect a variety of aquatic organisms (Fent et al., 2006). Therefore, it is very important to investigate the occurrence, distribution and mass balance of pharmaceuticals in WWTPs. However, most previous studies simply focused on the measurement of pharmaceuticals in raw influent and final effluent, as well as their overall removal efficiencies. Few studies considered the fate and distribution of pharmaceutical in individual treatment units along activated sludge treatment process, in order to comprehensively evaluate the biodegradation, persistence and partitioning behaviors in both aqueous and solid phases. In this work, the occurrence and removal efficiencies of the selected fifteen pharmaceuticals were investigated during the treatment processes in a WWTP located in East Lansing, Michigan. These target compounds were selected based on their common uses in the United States, and high frequency of detection in WWTPs (Kim et al., 2005; Karthikeyan and Meyer, 2006; Spongberg and Witter, 2008; Ding et al., 2011). Accelerated solvent extraction (ASE) and solid phase extraction (SPE) were combined to extract these pharmaceuticals from the wastewater and sludge samples collected from each unit of the WWTP, and analyzed by high-performance liquid chromatograph coupled to a tandem mass spectrometer (HPLC-MS/MS). Based on the data obtained, detailed mass balance analyses were performed to identify the effective elimination process and to estimate distribution for these pharmaceuticals in the treatment compartments. 2. Materials and methods 2.1. Materials and reagents The pharmaceuticals used in this study included TC, demeclocycline (DMC), chlortetracycline (CTC), oxytetracycline (OTC), doxycycline (DOC), meclocycline (MCC), sulfadiazine (SDZ), sulfamerazine (SMR), sulfamethazine (SMZ), SMX, tylosin (TYL), acetaminophen (AMP), ERY, LCM, CBZ and Caffeine (CAF). Their selected physicochemical properties are summarized in Supporting Information (SI) section SI-1. All the pharmaceuticals were purchased from Sigma–Aldrich Chemical Company (St. Louis, MO, USA) except DOC, which was purchased from Fisher Bioreagents (Pittsburgh, PA, USA). 13C6-sulfamethazine (13C6-SMZ) was obtained from Cambridge Isotope Laboratories (Andover, MA, USA). Stock solutions of reference pharmaceuticals were prepared in methanol and stored at 4 °C prior to use. Acetonitrile and methanol (HPLC-grade) were purchased from Sigma–Aldrich Chemical Company (St. Louis, MO, USA) and Mallinckrodt Baker Inc. (Phillipsburg, NJ, USA), respectively. Ultrapure water was obtained from a Milli-Q water purification system (Millipore, Billerica, MA, USA). 2.2. Sample collection and pretreatment Samples were collected in East Lansing Wastewater Treatment Facility in Michigan during May and December 2010. A simplified flow chart of the plant and all sampling sites are shown in section SI-2. Six batches of wastewater samples were taken, including three batches of 6 h and three batches of 24 h composite samples. One liter of each sample was collected in polypropylene bottles, and immediately put into ice-packed cooler, and transported to the laboratory. The wastewater samples were collected from raw influent, pretreatment effluent, primary effluent after the primary clarification, effluent after the aeration tank, secondary effluent

after the secondary clarification, and tertiary effluent after the post-filtration aeration. These samples were filtered through 0.45 lm nitrocellulose membrane (MF-MilliporeTM, Billerica, MA, USA) and transferred into pre-washed 1 L amber glass bottles within 10 h after collection. The samples were kept at 4 °C until they were extracted. Grab sludge samples were taken from the outlet of primary clarifiers (primary sludge), secondary sludge storage tanks (waste sludge), and outlet of the dewatering system (dewatered sludge). The collected primary sludge and waste sludge samples were immediately centrifuged at a speed of 10 000 rpm at 4 °C for 10 min, then the supernatants were decanted, and the solid fractions were frozen at 20 °C. Dewatered sludge was directly frozen without centrifugation. Samples were subsequently freeze-dried and finely grounded (<0.5 mm), and stored in amber glass bottles at 4 °C until extraction. 2.3. Sample extraction and cleanup For the preparation of wastewater samples, SPE was performed using 200 mg/6 mL Oasis HLB sorbent cartridges (Waters, Milford, MA, USA) connected to a 12-port PrepSep vacuum extraction manifold (Fisher Scientific, Fair Lawn, NJ, USA). The HLB cartridge was preconditioned with 3.0 mL of methanol, 1.0 mL of 0.1 M hydrochloric acid and 5.0 mL of Milli-Q water. One hundred milliliters of wastewater sample was percolated through the cartridge at a flow rate of approximately 3 mL min1 using the vacuum manifold. After percolation, the HLB cartridge was rinsed with 5.0 mL of MilliQ water and the eluent was discarded. Subsequently, the cartridge was eluted with 5.0 mL of methanol and water mixture (v/v = 75/ 25) containing 150 mg L1 EDTA. The eluate was collected, reduced to 1.0 mL by a gentle nitrogen flow at room temperature, and transferred into an amber autosampler vial for HPLC-MS/MS measurement. All samples were prepared in duplicate. For the sludge samples, an automated Dionex ASE 200 accelerated solvent extractor (Sunnyvale, CA, USA) equipped with a solvent controller was used for extraction. A mass of 500 mg of freeze-dried sludge sample was weighed to 22 mL stainless steel extraction cell with glass-fiber membranes at both ends, and thoroughly mixed with c.a. 5 g diatomaceous earth (inert matrix) to prevent aggregation. In our previous study (Ding et al., 2011), the extraction procedure applied here had been optimized for extraction solvent, extraction temperature and pressure with the target pharmaceuticals in sludges. Acetonitrile-water mixture (70/30, v/ v) was selected as extraction solvent with extraction temperature at 100 °C and pressure at 1500 psi. The preheating period, heating period and static extraction period were set as 5 min, 5 min and 15 min, respectively. The total flush volume was set as one cell volume, and the purge time was 120 s. The extraction was repeated with three cycles, and the collected extracts were transferred to a volumetric flasks, and diluted with Milli-Q water to 500 mL. The pH was adjusted to 3.0 using 40% sulfuric acid. Subsequently, 100 mL of the diluted sample was extracted and concentrated using the method described above for the preparation of wastewater sample. In this study, surrogate standards were added prior to ASE and SPE extraction. Specifically, five microliters of 236 mg L1 MCC was added as surrogate standard for tetracyclines (TCs), and 1 lL of 100 mg L1 13C6-SMZ for sulfonamides. Blank control consisting of diatomaceous earth was prepared. All samples were conducted in triplicate. 2.4. Analytical method Detailed analytical approach was described by Ding et al. (2011). The prepared samples were analyzed using a Shimadzu liquid chromatography system coupled to an Applied Biosystems API 3200 tri-

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ple quadrupole mass spectrometer (Foster City, CA, USA) equipped with an electrospray ionization source. The injection volume was 10 lL. Separation of pharmaceuticals was achieved using a Phenomenex Luna C18 column (150 mm  4.6 mm, 3 lm particle size and 100 Å pore size) (Torrance, CA, USA). The packing type of the column was end-capped. A binary gradient elution was used with a flow rate of 0.35 mL min1. Mobile phase A (water) and phase B (acetonitrile) were both acidified with 0.1% formic acid. The gradient was programmed. From 0 to 2 min phase A decreased from 100% to 88%, phase B correspondingly increased to 12%, and held it until 20 min. From 20 to 22 min phase B increased to 25%, and then kept on increasing to 40% at 35 min. From 35 to 40 min phase B increased to 80% and held it until 42 min. The mass spectrometer was performed at the positive mode with ion spray voltage of 5500 V and temperature of 600 °C. The declustering potentials were optimized for each pharmaceutical ranging from 17 to 63 V. Pharmaceutical concentration was quantified using multiple reaction monitoring (MRM) mode with the transition of the precursor to the most abundant product ion. These values and the detection limits are reported in Table 1. The reported concentrations of pharmaceuticals were estimated using the external standard method. The average extraction efficiencies of the whole procedure (n = 8) were 49% for CTC with standard deviation (SD) of 5.4%, 55% for DMC (SD = 5.0%), 68% for DOC (SD = 1.1%), 52% for OTC (SD = 5.3%), 54% for TC (SD = 8.7%), 64% for SDZ (SD = 5.2%), 95% for SMZ (SD = 2.7%), 71% for SMR (SD = 5.3%), 78% for SMX (SD = 8.0%), 79% for ERY (SD = 9.8%), 77% for TYL (SD = 5.4%), 84% for LCM (SD = 9.9%), 88% for CBZ (SD = 3.1%), 85% for AMP (SD = 10%), 80% for CAF (SD = 3.8%). The relatively small SDs (<10%) suggest the satisfactory repeatability of the method. Total organic carbon (TOC) contents for wastewater and sludge samples were determined using a Shimadzu TOC-VCPN analyzer equipped with solid sample module (SSM-5000A) for combustion. A XL20 pH-meter (Fisher Scientific, Pittsburge, PA, USA) with a glass pH electrode was used to measure wastewater pH values.

concentrations with the corresponding average flow along each wastewater treatment step. The equation can be expressed as:

maq ¼ Q aq C aq

ð1Þ

ms ¼ Q s C s

ð2Þ

where maq and ms (lg d1) are the mass flux of pharmaceutical calculated in aqueous and sludge phase, respectively. Qaq (L d1) and Qs (kg d1) are wastewater and sludge flow, respectively. Caq (lg L1) and Cs (lg kg1) are the average pharmaceutical concentrations measured in the wastewater and sludge, respectively. Normally, the removal of pharmaceuticals during municipal wastewater treatment process can be attributed to aerobic and anaerobic biotransformation, sorption to sludge and volatilization. In this study, the volatilization process was negligible because of the very low Henry constants of the investigated pharmaceuticals (SI-1). Accordingly, biodegradation and sorption are assumed to be the primary mechanisms for the removal of these substances. The mass flow calculation can be written as:

minf ¼ meff þ mbio þ msor

ð3Þ

where minf (kg d1) and meff (kg d1) are mass input and output of the treatment system. mbio (kg d1) and msor (kg d1) refer to the mass of pharmaceutical lost due to biodegradation and sorption, respectively. Here, mbio values were estimated by subtracting meff and msor from minf. The degree of elimination in percentage for biodegradation (Rbio, %) and sorption (Rsor, %) can be calculated using following equations.

mbio  100 minf msor ¼  100 minf

Rbio ¼

ð4Þ

Rsor

ð5Þ

Percentage of each compound (Ri, %) in effluents from different treatment units can be calculated based on the following equation. 2.5. Examination of mass balance

Ri ¼

The average mass flow of each detected pharmaceutical was calculated by multiplying the sum of aqueous and sludge phase

mi  100 minf

ð6Þ

where mi (kg d1) is mass flux in effluents from different units.

Table 1 Precursor and product ions, collision energy (CE), instrumental detection limit (IDL) and method detection limits (MDL) for the investigated pharmaceuticals. Pharmaceuticals

CTC DMC DOC OTC TC SDZ SMR SMZ SMX ERY TYL LCM CBZ AMP CAF a

Precursor ion [M + H]+ (m z1)a

479.2 465.0 445.4 461.3 445.4 250.9 264.9 279.0 254.1 734.6 916.5 407.3 237.1 152.0 195.0

Product ions (m z1)

462.2, 448.1, 428.2, 444.2, 428.2, 156.1, 172.4, 186.1, 156.2, 576.4, 771.9, 359.2, 194.3, 110.0, 138.0,

444.2 430.3 339.3 426.4 339.3 108.2 155.7 162.0 107.8 558.4 174.0 126.3 165.4 93.0 110.0

CE (V)

27 27 31 21 31 37 22 23 22 32 47 33 23 23 29

IDL (pg)

10.1 14.6 17.1 0.6 26.9 9.3 2.0 2.2 0.1 55.9 4.7 0.7 0.1 0.1 0.1

MDL Wastewater (ng L1)

Sludgeb (lg kg1)

LOD

LOQ

LOD

LOQ

27.6 38.3 45.2 6.5 68.2 26.7 10.2 9.9 2.3 86.5 13.3 5.4 2.2 2.6 3.1

92.0 127.7 150.7 21.7 227.3 89.0 34.0 33.0 7.7 288.3 44.3 18.0 7.3 8.7 10.3

37.2 25.9 4.6 13.7 146.0 15.0 5.5 0.6 1.0 6.6 5.4 0.8 2.9 30.7 8.4

124.0 86.3 15.3 45.7 486.7 50.0 18.3 2.0 3.3 22.0 18.0 2.7 9.7 102.3 28.0

Mass to charge ratio. Based on dry weight. LOD was determined as three times the standard deviation when injecting 10 lL of standards for 10 times. LOQ was determined as 10 times the standard deviation when injecting 10 lL of standards for 10 times. The procurer and two product ions were used to identify the pharmaceutical in the sample; the product ion in bold was used in the ion transition pair for quantification. b

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3. Results and discussion 3.1. WWTP operation The basic characteristics of wastewater collected in East Lansing WWTP are summarized in section SI-3. The average wastewater flow during the sampling periods was 4.54  104 m3 d1. The annual average dewatered sludge production was about 3300 tons; the solid content of the sludge was 19.6%. 3.2. Concentrations and removal of pharmaceuticals in wastewater The average pharmaceutical concentrations in wastewater and sludge samples collected from different sites in the WWTP are reported in section SI-4. DMC, SMR, ERY and TYL were not detected in raw influent during the sampling periods, while DOC, OTC, SDZ, SMX, LCM, CBZ, AMP and CAF were detected in all collected samples. CTC, TC and SMZ were detected at the levels above their limits of detection (LOD) in raw influent, however, none of them was found in final treated effluent (Fig. 1). SMX was detected in all wastewater samples with concentration as high as 1566 ng L1 in raw influent and 178 ng L1 in final effluent. Combination of SMX and trimethoprim was commonly used to treat methicillin resistant S. aureus (MRSA) infections due to the resistance to clindamycin (Genç et al., 2008). Such a high concentration detected in wastewater is probably due to its high consumption in clinical treatment. High concentration of SMX in sewage influent was also reported in WWTPs in other countries, such as 7190 ng L1 in China (Peng et al., 2006), 674 ng L1 in Sweden (Lindberg et al., 2005) and 820 ng L1 in Germany (Ternes et al., 2007). Several previous studies indicate that about 15% of SMX was excreted unchanged from human body (Hirsch et al., 1999), and approximately 50% was metabolized to N4-acetylsulfamethoxazole (N4-AcSMX) (Göbel et al., 2005). However, a similar structure product N4-acetylsulfamethazine (N4-AcSMZ) can transform back to SMZ during the storage of manure (Berger et al., 1986). The study by Göbel et al. (2005) also showed SMX present in wastewater could be derived from N4-AcSMX during the biological treatment process. It was observed that 63% of SMX was removed during the secondary biological treatment which could be attributed to potential biodegradation because sorption to the sludge was minimal (describe below). Göbel et al. (2005) also reported a median reduction of 68% for SMX and N4-AcSMX in biological treatment processes.

SDZ was detected in all wastewater samples above LOD, but at levels less than its limit of quantification (LOQ). SDZ concentration in the raw influent and the final effluent was 37 ng L1 and 27 ng L1, respectively. The removal efficiency of SDZ from the aqueous phase was about 27%, which was much lower than that reported at 50% by Xu et al. (2007). SMZ was only found in raw influent with the average concentration of 26 ng L1. Tetracyclines were frequently detected in the wastewater samples. DOC and OTC were found in all of the collected wastewater samples with the average concentrations of 738 ng L1 for DOC in the raw influent and 370 ng L1 in the final effluent. The elimination rate was estimated at about 50%, which was close to the results from Carballa et al. (2004) and Lindberg et al. (2005). For OTC, the average concentration was 29 ng L1 in the raw influent and 17 ng L1 in the final effluent. CTC and TC concentrations ranged between 178 and 309 ng L1 in the raw influent, whereas neither was detected in the final effluent. The greater elimination rate for tetracyclines in wastewater treatment could be due to their strong sorption to the sludge (Kim et al., 2005). CBZ was detected in all collected wastewater samples with the average concentrations of 110 ng L1 in the raw influent and 155 ng L1 in the final effluent. CBZ concentration in the raw influent was 41% less than that in the final treated effluent. This result suggests that the treatment process failed to effectively remove CBZ from wastewater. Spongberg and Witter (2008) reported that CBZ concentration in the effluent of an urban WWTP in Northwest Ohio increased two times compared with that in the influent. This could be attributed to the very inefficient treatment for CBZ removal in most conventional WWTPs, i.e., elimination rate of 0.1% and 7% (Ternes, 1998; Yamamoto et al., 2007). In addition, sludge-sorbed CBZ can be released into water phase when the sorption equilibrium shifts to aqueous phase as CBZ concentration decreases in wastewater. The average concentration of LCM was 58 ng L1 in the raw influent and 35 ng L1 in the final effluent. The corresponding elimination rate was about 40% in the aqueous phase, which was a little greater than those reported by Watkinson et al. (2007) and Behera et al. (2011). LCM is an organic base with a pKa = 7.6 and hence manifests predominantly as cationic species at pH 7. Inorganic cations from electrolytes present in wastewater compete with cationic LCM for negatively charged sorption sites on the sludge, hence reducing sorption and transformation (Wang et al., 2009; Chen et al., 2010). AMP and CAF were the two pharmaceuticals found in the raw influent at the concentration as high as 61 and

Fig. 1. Concentrations of the investigated pharmaceuticals detected in the raw influent and in the final effluent. Circles in the boxplots refer to the mean values (n = 6). Grey and white rectangular boxes represent the raw influent and the final effluent, respectively.

P. Gao et al. / Chemosphere 88 (2012) 17–24

41 lg L1. The treatment process appeared to effectively remove these two compounds from water, reducing their concentrations by three orders of magnitude down to 98 and 76 ng L1 for AMP and CAF in the final effluent. The corresponding removal efficiencies reached >99%.

3.3. Concentrations of pharmaceuticals in sludge Most of the investigated pharmaceuticals found in the aqueous phase were also detected in collected sludge samples. The associated concentrations of the detected pharmaceuticals in the dewatered sludge are presented in Fig. 2. SMX, which was frequently detected in wastewater, was also found in sludge samples with the concentrations of 34, 67 and 27 lg kg1 dw in the primary sludge, the waste sludge and the dewatered sludge, respectively. Other studies also found SMX in sludge of WWTPs with concentration of 5.7 and 68 lg kg1 dw (Göbel et al., 2005; Spongberg and Witter, 2008). SDZ was detected only once at the concentration > LOD in the collected sludge samples. The concentration was 39 lg kg1 dw in the waste sludge and 49 lg kg1 dw in the dewatered sludge. SMZ was also found in the waste sludge and the dewatered sludge but at relatively low concentrations in the range between 3.5 and 7.0 lg kg1 dw. Sulfonamide antibiotics generally present as neutral and anionic species at pH 7 (Le-Minh et al., 2010), and have low logKow values (SI-1). Therefore, sorption to the sludge is expected to be weak due to electrostatic repulsion from the negatively charged functional groups in the activated sludge. Tetracyclines generally manifest strong sorption to sludge via complexation with metals associated with sludge and cation exchange reaction. In the present study, the relatively higher concentration was observed for tetracyclines present in the sludge phase. For instance, DOC concentration was 762 lg kg1 dw in the primary sludge, 313 lg kg1 dw in the waste sludge and 568 lg kg1 dw in the dewatered sludge. TC was detected in the waste sludge and the dewatered sludge with the concentration of 750 and 566 lg kg1 dw, but was not found in the primary sludge. Kim et al. (2005) showed that the elimination of TC was positively related with sludge retention time (SRT). A significant TC reduction was observed when SRT increased from 3 to 10 d. Temperature, pH, Ca2+ and Mg2+ concentrations as well as dissolved organic matter content also affect TC sorption to sludge (Christian et al., 2003; Gu and Karthikeyan, 2008). OTC concentrations ranged between 17 and 18 lg kg1 dw, which were much less than DOC and TC

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concentration in sludge. This is due to the fact that the amount of OTC input to the WWTP was one order of magnitude less than DOC and TC (Table 2). CTC was found only once in the primary sludge with a concentration of 61 lg kg1 dw, and was not detected in the waste sludge and the dewatered sludge. CBZ is reported recalcitrant to biodegradation and abiotic transformation during wastewater treatment processes (Clara et al., 2005). In this study, CBZ was detected in all collected sludge samples, but with relatively low concentrations from 17 to 30 lg kg1 dw. This is consistent with previous studies. Miao et al. (2005) found 70 lg kg1 dw of CBZ in the untreated biosolids in a WWTP consisting of secondary sewage treatment processes. The mass of CBZ distribution in biosolids only accounted for 0.1% of the total loading in the sewage influent of the WWTP. The activated sludge treatment demonstrated a minimal removal of CBZ in the sludge (Clara et al., 2005). LCM concentration was 14 lg kg1 dw in the primary sludge, 41 lg kg1 dw in the waste sludge and 15 lg kg1 in the dewatered sludge. AMP was found in all sludge samples with average concentration of 109, 111 and 113 lg kg1 dw in the primary sludge, the waste sludge and dewatered sludge. CAF concentration was 73 lg kg1 dw in the dewatered sludge. In general, AMP and CAF appeared to be readily removed from the aqueous phase during the biological treatment process. More than 90% of AMP could be removed in activated sludge treatment by biotic transformations, while sorption to sludge played a negligible role (Jones et al., 2007; Behera et al., 2011).

3.4. Mass balance analysis The budget of pharmaceutical mass and distributions in water and sludge are estimated in order to further evaluate the operative removal process in the WWTP. The mass balance was examined for the pharmaceuticals detected in the collected samples. The mass flow was estimated using Eqs. (1) and (2), and the results are summarized in Table 2. All the pharmaceuticals demonstrated a relatively low percentage of retention in the dewatered sludge (Table 2, Fig. 3). TC was found to be apparently distributed in the sludge, but only 7.1% of the initial loading. The distributions of other compounds in the sludge phase were <5.2%. CTC, SMZ and SMX were all found in the raw influents, and had a very low retention in the dewatered sludge (less than 0.5% of the initial loading). The treatment process in the WWTP effectively removed nearly all mass loading of CTC

Fig. 2. Concentrations of the investigated pharmaceuticals detected in the dewatered sludge. Circles in the boxplots represent the mean values (n = 5).

100 NA 47 37 93 22 NA 99 >89 NA NA 39 -41 >99 >99 NA NA 3.0 2.2 7.1 5.2 NA 0.5 <0.1 NA NA 1.0 0.6 <0.01 <0.01 NA, not currently available. Mass flux was calculated according to Eqs. (1) and (2). Rbio and Rsor were calculated using Eqs. (4) and (5), respectively.

NA NA 1.0 ± 0.5 0.03 ± 0.02 1.0 ± 0.5 0.1 ± 0.03 NA 0.01 ± 0 0.1 ± 0.03 NA NA 0.03 ± 0.02 0.03 ± 0.02 0.2 ± 0.1 0.1 ± 0.05 NA NA 1.3 ± 1.3 0.1 ± 0.03 3.1 ± 1.7 0.2 ± 0.2 NA 0.03 ± 0 0.3 ± 0.2 NA NA 0.2 ± 0.04 0.1 ± 0.05 0.5 ± 0.3 0.4 ± 0.6 0.2 NA 2.3 ± 1.5 0.1 ± 0.01 NA 0.2 ± 0.04 NA NA 0.1 ± 0.1 NA NA 0.04 ± 0.05 0.1 ± 0.1 0.3 ± 0.1 0.3 ± 0.1 NA NA 17 ± 17 0.8 ± 0.2 NA 1.2 ± 0.6 NA NA 8.1 ± 6.9 NA NA 1.6 ± 0.5 7.0 ± 2.5 4.5 ± 3.9 3.4 ± 2.7 NA NA 15 ± 18 0.6 ± 0.4 NA 1.3 ± 0.1 NA NA 22 ± 3.9 NA NA 2.6 ± 2.0 7.1 ± 2.2 3.2 ± 1.6 3.3 ± 2.1 2.4 ± 0.9 NA 26 ± 25 0.9 ± 0.2 7.0 ± 3.9 2.0 ± 0.3 NA NA 75 ± 32 NA NA 2.6 ± 3.5 7.7 ± 2.7 145 ± 134 138 ± 205 7.7 ± 1.8 NA 22 ± 16 0.9 ± 0.6 5.7 ± 2.0 1.6 ± 0.4 NA NA 58 ± 24 NA NA 6.1 ± 7.6 5.4 ± 1.2 2644 ± 1099 1737 ± 538 CTC DMC DOC OTC TC SDZ SMR SMZ SMX ERY TYL LCM CBZ AMP CAF

8.1 ± 8.2 NA 34 ± 25 1.3 ± 0.8 14 ± 4.2 1.7 ± 0.5 NA 1.2 ± 0.1 71 ± 26 NA NA 2.6 ± 3.6 5.0 ± 1.2 2800 ± 1493 1871 ± 550

5.5 ± 1.1 NA 20 ± 16 1.2 ± 0.3 7.0 ± 2.4 1.6 ± 0.3 NA NA 59 ± 23 NA NA 3.4 ± 5.1 6.0 ± 1.8 2561 ± 1238 2436 ± 794

Dewatered sludge Waste sludge Primary sludge Final effluent Secondary effluent Aeration effluent Primary effluent Pretreatment effluent Raw influent

Mass flux (g d1) Pharmaceuticals

Table 2 Mass flux of the investigated pharmaceuticals at different treatment units.

Fig. 3. Mass distribution of the detected pharmaceuticals in the WWTP. The greycolored bar represents the mass fraction in the final effluent, the black-colored bar represents the fraction in the dewatered sludge, and the white bar represents the loss of pharmaceuticals due to biodegradation.

NA NA 50 61 NA 73 NA NA 11 NA NA 60 141 <0.2 <0.2

Mass in effluent (%)

Mass lost in WWTP, Rbio (%)

P. Gao et al. / Chemosphere 88 (2012) 17–24

Mass in dewatered sludge, Rsor (%)

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and SMR (elimination rate >99%). For SMX, the biological treatment process contributed 89% removal of the loading, and 11% of SMX survived the treatment processes, and remained in the treated effluent. Mass balance analyses indicate that DOC, OTC, SDZ and LCM behaved similarly in terms of distribution in wastewater and sludge, as well as the loss due to biological treatment of the WWTP (Fig. 3). The distribution fractions in the final effluent ranged from 50% (DOC) to 73% (SDZ), and the lost fractions ranged from 22% (SDZ) to 47% (DOC). It was noted that CBZ mass in the treated effluent was 1.4 times the mass in the influent to the WWTP (Fig. 3). There was only 0.6% of CBZ in the dewatered sludge. These results indicate that the conventional wastewater treatment processes can not eliminate CBZ in the influent into WWTPs (Miao et al., 2005). In contrast, the treatment processes could effectively remove (nearly 100%) AMP and CAF in the influent, and the contribution of sorption process to the removal can be negligible (Fig. 3). The distribution of the detected pharmaceuticals along the treatment units is shown in Fig. 4. The results show that during the pretreatment step DOC, OTC and TC decreased by 35%, 28% and 59%, while other pharmaceuticals remained nearly same. The primary treatment removed 27% for CTC, and no obvious reduction was observed for other pharmaceuticals. Significant reduction (>46%) was observed for many pharmaceuticals during the biological treatments (aeration tank and biological treatment), though the elimination was not apparent to DOC, SDZ, LCM and CBZ. The removal percentages were approximately 15% and 18% for DOC and SDZ. No reduction was found for CBZ and LCM. The tertiary treatment processes including chlorine disinfection and sand filtration significantly diminished SMX and LCM by 19% and 38% from the effluents of the biological treatment. No apparent pattern was found for the degradation of pharmaceuticals in the sludge due possibly to the lower fractions in the solid phase.

4. Conclusions Eleven out of the fifteen investigated pharmaceuticals were detected in sewage influent of the WWTP (East Lansing, Michigan) employing conventional biological treatment process. Most of the pharmaceuticals found in the wastewater were also detected in the sludge phase. Among the investigated pharmaceuticals, AMP and CAF were found in the raw influent at high concentrations 62 ± 33 and 41 ± 12 lg L1. Removal rates varied between 22%

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Fig. 4. Percentage (Ri, %, calculated applying Eq. (6)) of detected pharmaceuticals along the WWTP.

(for SDZ) to nearly 100% (for CTC) during the treatment process. CBZ was persistent throughout the conventional treatment, and demonstrated minimal degradation. The mass balance analyses indicate that biodegradation is the major process responsible for the removal of pharmaceuticals, while the contribution of sorption is insignificant. Overall, the results from this study improve the understandings of fate and transport of pharmaceuticals in wastewater treatment facilities, and help assess the potential risks posed to surface water quality.

Acknowledgements We gratefully acknowledge the staff from in East Lansing WWTP for providing help on sample collection, as well as the information needed in this study. Also, we would like to thank Mr. Borislav Lazarov for his assistance with experiments.

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