Occurrence and fate of tetracycline and degradation ...

Environ Monit Assess (2015) 187:750 DOI 10.1007/s10661-015-4978-4

Occurrence and fate of tetracycline and degradation products in municipal biological wastewater treatment plant and transport of them in surface water Murat Topal & E. Işıl Arslan Topal

Received: 15 May 2015 / Accepted: 9 November 2015 # Springer International Publishing Switzerland 2015

Abstract The aims of this study are to investigate the fate of tetracycline (TC) and degradation products (DPs) in municipal biological wastewater treatment plant (MBWWTP) located in Elazığ City (Turkey) and to determine the occurrence and transport of TC and DPs in surface water (SW) (Kehli Stream) which the effluents of the plant discharged. The aqueous phase removal of TC, 4-epitetracycline (ETC), 4-epianhydrotetracycline (EATC), and anhydrotetracycline (ATC) in the studied treatment plant was 39.4±1.9, 31.8±1.5, 15.1±0.7, and 16.9±0.8 %, respectively. According to the analyses’ results of SW samples taken from downstream at every 500-m distance, TC and DPs decreased by the increase in the distance. In downstream, at 2000 m, TC, ETC, EATC, and ATC were 4.12±0.20, 6.70±0.33, 8.31± 0.41, and 3.57±0.17 μg/L, respectively. As a result, antibiotic pollution in the SW that takes the effluent of MBWWTP exists. Keywords Tetracycline . Surface water . Micropollutants . Antibiotic . Sludge . Turkey Electronic supplementary material The online version of this article (doi:10.1007/s10661-015-4978-4) contains supplementary material, which is available to authorized users. M. Topal (*) General Directorate of State Hydraulic Works, 9th District Office, Elazığ, Turkey e-mail: [email protected] E. I. Arslan Topal Faculty of Engineering, Department of Environmental Engineering, University of Firat, Elazığ, Turkey

Introduction Many active antibiotics are not completely metabolized during therapeutic use and thus enter sewage through excretion in an unchanged form (Le-Minh et al. 2010). The intentional disposal of unused drugs into the sewer (Kummerer 2003), discharges from veterinary clinics, and runoff from agricultural applications into municipal sewers also contribute to the quantities of antibiotics found in sewage (Le-Minh et al. 2010). Unfortunately, municipal wastewater treatment plants (MWWTPs) are generally unable to effectively remove either unaltered or metabolized forms of pharmaceuticals from wastewaters (WWs) (Al-Aukidy et al. 2012). Because of the observed concentrations of pharmaceuticals and personal care products (PPCPs) in raw WWs and limited effectiveness of secondary treatments, MWWTP effluents are the main introduction pathway for PPCPs into environment (Urtiaga et al. 2013). Antibiotics in surface waters (SWs) have the potential to disrupt key bacterial cycles/processes critical to aquatic ecology (nitrification/denitrification) or agriculture (soil fertility) and animal production (Watkinson et al. 2009; Kummerer 2010). Tetracyclines (TCs) are broad-spectrum antibiotics used in human and veterinary medicine for the treatment of infectious diseases. TCs are poorly absorbed and are excreted via feces and urine in their original form and/or as primary metabolites. They get into the WWs and proceed to WWTPs (Škrášková et al. 2013).

750

Environ Monit Assess (2015) 187:750

Page 2 of 9

TC was detected in various environments. Luo et al. (2011) reported TC in the Haihe River (China) around 26 ng/L (Lian et al. 2013). Kim and Carlson (2007) determined TC at 20 ng/L in the Cache La Poudre River, USA. In the USA, residues of TCs have been detected at 0.11 μg/L levels in SWs, while in sewage treatment plants (STPs), concentrations of 0.52 μg/L in influents have been reported (Kolpin et al. 2002; Karthikeyan and Meyer 2006; Jeong et al. 2010). High proportion of TC with different forms would get rich in groundwater, SW, and soil due to their ineffective biodegradability (Gao et al. 2012; He et al. 2013). TCs have high risk of occurrence of antibioticresistant bacteria in activated sludge, and TC-resistant genes have been found at outfall of or downstream MWWTPs (Saitoh et al. 2014). Szczepanowski et al. (2009) measured 23 TC resistance genes in activated sludge and effluent from WWTP. Auerbach et al. (2007) and Mispagel and Gray (2005) reported occurrence of TC resistance genes in small- and large-scale WWTPs (Hopkins and Blaney 2014). Under photochemical degradation, TCs will form a number of degradates, many of which are highly soluble and have been shown to be more stable form in receiving waters (Watkinson et al. 2009). Degradation products (DPs) can be as active and/or toxic as their parent (Watkinson et al. 2007). In order to reduce negative impacts on environment and human health, it is necessary to understand input sources (e.g., WWTP effluents) for antibiotics (Zhou et al. 2013a). TCs are one of the most frequently detectable antibiotics in SW resources that receive discharges from agricultural effluents and MWWTPs (Saitoh et al. 2014). Because of its high risk, treatment of TC in effluents is important. Also, determination of occurrence of TC in WWTP sludges applied to agricultural lands is important because of the risk for food chain. In this study, fate of TC and DPs in Elazığ MBWWTP and their occurrence and transport in SW (Kehli Stream) which takes effluent of the plant were investigated. There is a significant lack of study concerning the occurrence and transport of the antibiotics in SWs in Turkey. In Turkey, to the best of our knowledge, this is the first study which deals with the occurrence and transport of the TC and DPs in the SW. This study is also important because of its contribution to the literature about the occurrence and transport of TC in various environments of world.

Materials and methods Chemicals In the study, TC was purchased from Sigma-Aldrich while the DPs (4-epitetracycline (ETC), 4epianhydrotetracycline (EATC), and anhydrotetracycline (ATC)) were purchased from Acros Organics (New Jersey, USA). The other chemicals used were methanol (Carlo Erba), methylene chloride (Fisher Chemical), acetonitrile, and formic acid (J.T. Baker). They were all of HPLC grade. Hydrochloric acid (HCl) (J.T. Baker), sodium hydroxide (NaOH) (Acros Organics), ammonia solution (NH3·H2O) (Carlo Erba), and ethylenediamine tetraacetic acid disodium (Na2EDTA) (Sigma-Aldrich) were all of analytical reagent grade. Oasis HLB (500 mg, 6 cm3) and Oasis MAX (60 mg, 3 cm3) cartridges were purchased from Waters Corporation (Milford, MA, USA). The ultrapure water used through the study was supplied by a Zeneer power water purification system. Sample collection WW samples were obtained from Elazığ MBWWTP that serves a population of 383,975. Hydraulic retention times in various stages of MBWWTP were taken into account when sampling WW influents and effluents to reduce error. MBWWTP was specifically chosen for its location because effluent of it is discharged to a SW and finally reaches to a dam lake (Keban Dam Lake) which the waters of it are used for various purposes. Waters of the lake are used as drinking water in rural areas and also as irrigation water for many agricultural activities. Furthermore, the lake is one of the important water resources in Turkey because of the fish hatchery. Therefore, the pollution that SWs transport to the lake is important (Topal and Arslan Topal 2014). Residential, public, commercial, hospital, and industrial WWs enter to MBWWTP via sewage system. Plant treats between ≈36,000 and 54,000m3/day of WW according to the season (Topal et al. 2015). The location of MBWWTP is given in Fig. 1. MBWWTP was put into operation in 1994 and revised in 2007. Plant was projected according to the conventional activated sludge system. It is operated with three fundamental mechanisms (pretreatment, biological treatment, and sludge disposal). Flow diagram is


Page 3 of 9 750

Fig. 1 The location of Elazığ MBWWTP

given in Fig. S1. WWs are treated in plant and then discharged into Kehli Stream (Topal and Arslan Topal 2011; Topal et al. 2012). MBWWTP has ≈28–42 % flow contribution to stream. WWs discharged from MBWWTP flow to Keban Dam Lake from the stream. The distance from the discharge of the effluent of MBWWTP to the lake is ≈4 km. WW samples in MBWWTP were taken from influent of MBWWTP, effluent of bar screen and grit chamber, effluent of primary clarifier, effluent of aeration tank, and effluent of secondary clarifier (effluent of MBWWTP). WW samples were collected as 24-h composite samples. Sludge samples were taken from sludge drying beds of MBWWTP. SW samples were taken from Kehli Stream that takes effluent of MBWWTP. Kehli Stream is one of the important SW resources of Elazığ City. According to the calculations done by at-site flood frequency analysis, total precipitation area is 223.9 km2 and flows of Q100 and Q500 are 441 and 592 m3/s, respectively (Topal and Arslan Topal 2014). It has yearly average flow rate of ≈1.5 m3/s. It flows from south to north (Fig. 1). Agricultural lands, sheds, some little corporations, and residentials exist along it. SW sampling points were at different distances in Kehli Stream (at 500 m from upstream of discharge location, at the point of WW effluent discharge, at 500,

1000, 1500, and 2000 m from downstream of discharge location). All sampling points are shown in Fig. S2 as schema. Samples from stream were collected approximately 0.1 m below the surface along a cross section of the site. SW samples were collected as grab samples. Samples were taken from nine different points at each sampling point, and they were mixed and the sample analyzed were taken from that mixed sample. The different collection time points were adjusted for flow in stream so that we sampled the same body of water as we moved downstream (Fig. S3). WW and SW samples were collected in 1-L bottles, which had previously been washed with 0.5 g/L Na2EDTA water solution, followed by methanol, and then rinsed with ultrapure water and dried before use. Collected samples were stored on ice in a dark container during transportation to the laboratory and processed immediately on the same day. The triplicate samples were analyzed. Sample extraction The method reported by Jia et al. (2009) was used for extraction of WW and SW samples. The samples collected were filtered with glass microfiber filter (0.7 μm, Whatman). A 150-mL sample was added with 0.5 g/L Na2EDTA and acidified to pH 3.0 with HCl. Oasis HLB cartridges were preconditioned with 6 mL of methylene chloride, 6 mL of methanol, and 6 mL of ultrapure water

750

Page 4 of 9

containing 0.5 g/L Na2EDTA (adjusted to pH 3.0 with HCl). Samples were passed through cartridges at a flow rate of 3 mL/min. After being rinsed with 10 mL of ultrapure water, cartridges were dried under flow of nitrogen. Dried cartridges were eluted with 6 mL of methanol. After eluates were collected in an amber vial, they were dried under a gentle flow of nitrogen. Then, they were reconstituted to 0.3 mL with methanol. After extracts were diluted to 8 mL by ultrapure water (adjusted to pH 7.0 with 5 % NH3·H2O), the solutions were then applied to the Oasis MAX cartridges which had been conditioned with 1 mL of methanol, 1 mL of 5 N NaOH, and 1 mL of ultrapure water. Cartridges were rinsed with 1 mL of 5 % NH3·H2O, followed by 1 mL of methanol. Elution was performed with 3 mL of acetonitrile/water containing 1 % formic acid (50/50, v/v) mixed reagents. Extracts were concentrated to 1.5 mL under stream of nitrogen and measured with ultrafast liquid chromatography-tandem mass spectrometry (UFLC–MS/MS) soon after they were prepared (Fig. 2). The extraction of the sludge samples taken from the MBWWTP was performed by the method used by Lillenberg et al. (2010). Two hundred fifty milligrams of dried sludge was extracted with 10 mL of 1:1 (v/v) mixture of acetonitrile and 1 % acetic acid, then homogenized with laboratory homogenizer DIAX 900 (Heidolph Instruments, Germany) 25000 rpm, sonicated (5′), vortexed (1′), and centrifuged at 8000 rpm. The supernatant was then separated and dried by nitrogen stream. Approximately 15 mL of 1 % acetic acid was added to the 1 mL of evaporation residue. The triplicate samples were extracted. UFLC–MS/MS instrumental analysis Concentrations of TC and DPs were analyzed using UFLC–MS/MS (Shimadzu UFLC coupled to 3200

Fig. 2 Sample extraction stages


QTRAP, Applied Biosystems). Separation of TC and DPs was achieved with a Waters ACQUITY UPLC BEH C18 column (1.7 μm; 2.1×100 mm). The injection volume was 10 μL (full loop). Acetonitrile (A) and ultrapure water containing 0.1 % formic acid (v/v) (B) were used as mobile phases. The gradient was as follows: The initial 10 % A was increased linearly to 20 % in 5 min, a further 20 % A was increased to 90 % over 4 min and kept for 0.5 min, and followed by an increase to 100 % A and held for 1 min. Finally, the gradient was returned to the initial conditions of 10 % A and held for 2 min to allow for equilibration. The flow rate was 0.2 mL/min. The column was maintained at 30 °C, and the sample room temperature was 20 °C. Mass spectrometry was performed using a AB Applied Biosystems (triple-quadrupole) detector equipped with an electrospray ionization.

MBWWTP removal efficiency calculations Antibiotic removal efficiency of MBWWTP was calculated by the following equation; Removal EfficiencyðREÞð%Þ ¼

Ci−Ce 100 Ci

ð1Þ

where Ci is antibiotic concentration of influent (μg/L) and Ce is antibiotic concentration of effluent (μg/L). Method validation Calibration curves were constructed from 1 to 100 μg/L. The concentration gradient was at 1, 2, 3, 4, 5, 10, 30, 50, and 100 μg/L, and mean coefficients of determination (R2) were 0.9965, 0.9996, 0.9850, and 0.9761 for TC, ETC, EATC, and ATC, respectively. The limit of detection (LOD) was 0.28, 0.30, 0.40, and 0.59 μg/L for TC, ETC, EATC, and ATC, respectively. The limit of


quantification (LOQ) values were 0.92, 0.99, 1.32, and 1.95 μg/L for TC, ETC, EATC, and ATC, respectively. Statistical analyses Experimental results were analyzed using the IBM SPSS Statistics 21 programme (USA), and values shown were the means of three replicates. Each point is the mean of three replicates. Error bars indicate the standard deviation.

Results and discussion In the study, firstly, TC and DPs were analyzed in WW those taken from different units of MBWWTP. TC is one of the most frequently detected antibiotics in WWTPs. TC in influent of MBWWTP was 10.3± 0.5 μg/L. Possible reason for this high value of TC is the usage habit. Turkey is the premier in 40 Europe countries in terms of antibiotic usage. ETC, EATC, and ATC concentrations in influent of MBWWTP were 12.2± 0.6, 9.78±0.48, and 6.35±0.31 μg/L, respectively. In the scientific literature, to the best of our knowledge, there is not any study which deals with the concentrations of ETC, EATC, and ATC in WWs. In effluent of bar screen and grit chamber, TC, ETC, EATC, and ATC concentrations decreased. TC, ETC, EATC, and ATC removals were 42.5±2.1, 36.4±1.8, 14.6±0.7, and 0.3±0.01 % in bar screen and grit chamber, respectively. Removals followed that order: TC> ETC>EATC>ATC. TC, ETC, EATC, and ATC in influent of primary clarifier which is the last step of primary treatment unit of MBWWTP were 5.92±0.29, 7.76±0.38, 8.35±0.41, and 6.33±0.31 μg/L, respectively. TC and ATC removals of primary clarifier were 9.8±0.4 and 8.2± 0.4 %, respectively. Removal depends mostly on sorption potential to suspended solids deposited from primary sedimentation (Leung et al. 2012). There was not any removal for ETC and EATC in primary clarifier. Their concentrations increased to 8.17 ± 0.40 and 9.57 ± 0.47 μg/L, respectively. This situation could be due to the limited stability of TCs in aqueous media (HallingSørensen et al. 2002). The abiotic degradation products or reversible epimers may be formed through hydrolysis or photolysis, including epi-tetracyclines and anhydrotetracyclines (Brain et al. 2005). The negative removal has also been ascribed to the daily concentration

Page 5 of 9 750

fluctuations during the sampling period or desorption of molecules from sludge and suspended particulate matter (Luo et al. 2014). TC, ETC, EATC, and ATC in influent of aeration tank were 5.34±0.26, 8.17±0.40, 9.57±0.47, and 5.81± 0.29 μg/L, respectively. There was removal for EATC and ATC in aeration tank. EATC and ATC removals of aeration tank were 11.2±0.5 and 4±0.2 %, respectively. According to the analyses’ results of effluent of aeration tank, TC and ETC increased. There was not any removal for TC and ETC. When negative removal percentages (concentration for an antibiotic in effluent was higher than in influent) are seen, this might be explained by the presence of antibiotic conjugates and/or metabolites that are reverted back during treatment into its original form and slow biological degradation of the antibiotic (Zhou et al. 2013a). Probably degradation of TC in aeration tank was slow in MBWWTP. In addition, some pharmaceuticals excreted with feces are probably partly enclosed in feces particles and released during biological treatment (Köck-Schulmeyer et al. 2013). This situation was probably seen for TC in our study. TC, ETC, EATC, and ATC concentrations in influent of secondary clarifier were 7.12±0.35, 9.01±0.45, 8.50 ±0.42, and 5.58±0.27 μg/L, respectively. TC, ETC, EATC, and ATC removals of secondary clarifier were calculated as 12.4±0.6, 7.7±0.3, 2.4±0.1, and 5.4± 0.2 %, respectively. Removals in secondary clarifier followed that order: TC>ETC>ATC>EATC. TC in effluent of MBWWTP was determined as 6.24 ±0.31 μg/L, in our study. The possible reason for high concentrations of TC in effluent of MBWWTP is high concentrations of TC in influent of Elazığ MBWWTP due to the usage habitat aforementioned. Removal efficiencies of TC and DPs in MBWWTP are given in Fig. 3. TC, ETC, EATC, and ATC removals of MBWWTP were 39.4±1.9, 31.8±1.5, 15.1±0.7, and 16.9±0.9 %, respectively. Removals followed that order: TC>ETC> ATC>EATC. In our study, TC and DPs in MBWWTP were not efficiently removed by activated sludge process. Biological treatment (e.g., activated sludge) often fails to degrade antibiotics (Chen et al. 2013). Therefore, tertiary and advanced treatment methods (membrane filtration, adsorption, natural treatment, wetland systems, etc.) should be used for removal of TC and DPs. TC, ETC, ATC, and EATC concentrations in sludges were 43.2±2.1, 49.6±2.4, 41.7±2.0, and 34.6±1.7 μg/ kg, respectively. TCs sorp onto sludge during treatment.


Page 6 of 9

Bar screen and grit chamber

Elaz ğ MBWWTP 100 80 60 40 20 0

50

Removal Efficiency (%)

Fig. 3 Removal efficiencies of TC and DPs in Elazığ MBWWTP


750

TC

ETC

EATC

40 30 20 10 0

ATC

TC

ETC



10 8 6 4 2 TC

ETC

EATC

ATC

Aeration tank

Primary clarifier 12

0

EATC

TC and DPs

TC and DPs

ATC

14 12 10 8 6 4 2 0

TC

ETC

EATC

ATC

TC and DPs

TC and DPs


Secondary clarifier 14 12 10 8 6 4 2 0

TC

ETC

EATC

ATC

TC and DPs

TCs are reported to interact strongly with clay, nature organic matter and metal oxides by cation exchange, surface complexation/cation, bridging hydrophobic partitioning, and electron donor–acceptor interactions (Zhou et al. 2013a). TCs are among the most abundant pharmaceuticals in sludges. The potential long-term effects of TC compounds and their continuous spread in environment make it necessary to monitor them in sludges as this information is critical to assess the exposure and potential risks (Pamreddy et al. 2013). Kehli Stream which takes the discharge of effluent of Elazığ MBWWTP flows to Keban Dam Lake. The waters of the lake are used for the purposes of drinking and usage. Agricultural lands exist close to stream. Water from stream is taken by the farmers for the purpose of irrigation of the agricultural lands. Therefore, the occurrence of the antibiotic residues in the SWs is important.

Concentrations of TC and DPs in WW (effluent of Elazığ MBWWTP) and SW (Kehli Stream) are given in Table 1. TC, ETC, EATC, and ATC were determined at levels of micrograms per liter (2.17±0.10, 3.57±0.17, 6.80± 0.34, and 4.44±0.22 μg/L, respectively) in SW samples taken from the upstream of MBWWTP effluent discharge location. This situation is probably because of the chicken farms, agriculture, and livestock activities in the lands close to Kehli Stream before the discharge of the effluent of MBWWTP into stream. Various antibiotics including TCs are used as feed additives in livestock industries (Zhou et al. 2013a). In livestock farms, most of WW and manure are directly/indirectly discharged into SW or applied onto agricultural land, resulting in the contamination of antibiotics in aquatic and terrestrial environments (Zhou et al. 2013b). TC sourced from the places aforementioned reaches to the

Environ Monit Assess (2015) 187:750 Table 1 Concentrations of TC and the DPs in MBWWTP and SW

Page 7 of 9 750

No.

Sampling point

1

MBWWTP effluent

6.24

8.32

8.30

5.28

2

Kehli Stream

2.17

3.57

6.80

4.44

3

Mixture point

8.30

9.21

8.81

5.53

4

After discharge (500 m)

6.23

8.87

8.80

5.53

5


6.18

7.50

8.77

5.43

6


5.03

7.03

8.72

3.85

7


4.12

6.70

8.31

3.57

Kehli Stream via different ways. Treated WW and sludge are used by farmers for agriculture in the studied location. TCs can enter SW via runoff from sludges and manure applied to agricultural lands. Therefore, it is not surprising to find TC and DPs in upstream of discharge location. TC and DPs in downstream of discharge location increased. This shows the importance of the removal efficiency of MBWWTP. Concentrations of TC and DPs (8.30±0.41, 9.21±0.46, 8.81±0.44, and 5.53± 0.27 μg/L, respectively for TC, ETC, EATC, and ATC) at the point of WW effluent discharge was higher than the values determined at the other sampling points of stream. In our study, generally, the most frequently detected compounds in MBWWTP effluent were also the most frequently detected ones in downstream of discharge location. In our study, concentrations of TC and DPs decreased with distance increase from the point of WW effluent discharge. TC and the DPs were seen in Kehli Stream at a distance of 2 km. In our study, concentrations of TC and DPs in downstream of discharge point were not decreased much (concentration of 8.30±0.41 μg/L at the point of WW effluent discharge decreased to 6.23±0.31 μg/L at a distance of 500 m from the point of discharge). The MBWWTP has ≈28±1.4–42±2.1 % flow contribution to Kehli Stream. Therefore, the situation is probably due to the low dilution of TC and DPs in stream. In terms of TC concentration, effluents of MBWWTP add a pollution load of 65.2±3.2 % into Kehli Stream. TC concentration of Kehli Stream increased 73.9±3.6 % after the discharge of the effluents of MBWWTP and increased from 2.17±0.10 to 8.30±0.41 μg/L. This pollution decreased at percentages of 24.9±1.2, 25.5±1.2, 39.4±1.9, and 50.4±2.5 after 500-, 1000-, 1500-, and 2000-m distances from the point of WW effluent discharge, respectively. TC was 4.12±0.20 μg/L at distance of 2000 m from the point of discharge.

TC (μg/L)

ETC (μg/L)

EATC (μg/L)

ATC (μg/L)

MBWWTP effluents transport pollution load of 57.1 ±2.8 % into stream in terms of ETC. ETC concentration of stream increased at a percentage of 61.24±3.06 because of discharge and increased from 3.57±0.17 to 9.21±0.46 μg/L. This pollution decreased at percentages of 3.7±0.1, 18.6±0.9, 23.7±1.1, and 27.3±1.3 after the point of discharge (500, 1000, 1500, and 2000 m), respectively. ETC was 6.70±0.33 μg/L at distance of 2000 m. MBWWTP effluents transport pollution load of 18.1 ±0.9 % into stream in terms of EATC. EATC concentration of stream increased at a percentage of 22.81± 1.14 because of discharge and increased from 6.80± 0.34 to 8.81±0.44 μg/L. This pollution decreased at percentages of 0.1±0.005, 0.5±0.025, 1.0±0.05, and 5.7±0.2 after the point of discharge (500, 1000, 1500, and 2000 m), respectively. EATC was 8.31±0.41 μg/L at distance of 2000 m. Effluents of MBWWTP transport pollution load of 15.9±0.7 % in terms of ATC into stream. Concentration of ATC in stream increased 19.7±0.9 % because of discharge of effluents and increased from 4.44±0.22 to 5.53±0.27 μg/L. This pollution did not change at the 500 m after the point of discharge and determined as 5.53±0.27 μg/L. ATC decreased 1.8±0.09, 30.4±1.5, and 35.4±1.7 % after 1000-, 1500-, and 2000-m distances from the point of discharge, respectively. ATC was 3.57±0.17 μg/L at distance of 2000 m.

Conclusions TC and DPs were at concentrations of micrograms per liter and micrograms per kilogram in MBWWTP effluents and sludge, respectively. These levels are important because both the treated WW and the sludges are used by farmers in agricultural soils. Concentrations of TC

750

Page 8 of 9

and DPs in SW increased after discharge of effluents of MBWWTP into it. Because of the growing concern to the environmental risks associated with the presence of antibiotics in various environments (SW, soil, e.g.), addition of the data obtained by this study to the current understanding of the occurrence and fate of TC and DPs in the environment is important.

References Al-Aukidy, M., Verlicchi, P., Jelic, A., Petrovic, M., & Barcelò, D. (2012). Monitoring release of pharmaceutical compounds: occurrence and environmental risk assessment of two WWTP effluents and their receiving bodies in the Po Valley, Italy. Science of the Total Environment, 438, 15–25. Auerbach, E. A., Seyfried, E. E., & McMahon, K. D. (2007). Tetracycline resistance genes in activated sludge wastewater treatment plants. Water Research, 41, 1143–1151. Brain, R. A., Wilson, C. J., Johnson, D. J., Sanderson, H., Bestari, K., Hanson, M. L., Sibley, P. K., & Solomon, K. R. (2005). Effects of a mixture of tetracyclines to Lemna gibba and Myriophyllum sibiricum evaluated in aquatic microcosms. Environmental Pollution, 138(3), 425–442. Chen, Y., Yu, G., Cao, Q., Zhang, H., Lin, Q., & Hong, Y. (2013). Occurrence and environmental implications of pharmaceuticals in Chinese municipal sewage sludge. Chemosphere, 93(9), 1765–1772. Gao, P., Munir, M., & Xagoraraki, I. (2012). Correlation of tetracycline and sulfonamide antibiotics with corresponding resistance genes and resistant bacteria in a conventional municipal wastewater treatment plant. Science of the Total Environment, 421–422, 173–183. Halling Sørensen, B., Sengelov, G., & Tjornelund, J. (2002). Toxicity of tetracyclines and tetracycline degradation products to environmentally relevant bacteria, including selected tetracycline-resistant bacteria. Archives of Environmental Contamination and Toxicology, 42(3), 263–271. He, Y., Chen, W., Zheng, X., Wang, X., & Huang, X. (2013). Fate and removal of typical pharmaceuticals and personal care products by three different treatment processes. The Science of the Total Environment, 447, 248–254. Hopkins, Z. R., & Blaney, L. (2014). A novel approach to modeling the reaction kinetics of tetracycline antibiotics with aqueous ozone. Science of the Total Environment, 468–469, 337– 344. Jeong, J., Song, W., Cooper, W. J., Jung, J., & Greaves, J. (2010). Degradation of tetracycline antibiotics: mechanisms and kinetic studies for advanced oxidation/reduction processes. Chemosphere, 78(5), 533–40. Jia, A., Xiao, Y., Hu, J., Asami, M., & Kunikane, S. (2009). Simultaneous determination of tetracyclines and their degradation products in environmental waters by liquid chromatography–electrospray tandem mass spectrometry. Journal of Chromatography A, 1216(22), 4655–4662.

Environ Monit Assess (2015) 187:750 Karthikeyan, K. G., & Meyer, M. T. (2006). Occurrence of antibiotics in wastewater treatment facilities in Wisconsin, USA. Science of the Total Environment, 361(1–3), 196–207. Kim, S. C., & Carlson, K. (2007). Temporal and spatial trends in the occurrence of human and veterinary antibiotics in aqueous and river sediment matrices. Environmental Science and Technology, 41, 50–57. Köck-Schulmeyer, M., Villagrasa, M., López de Alda, M., Céspedes-Sánchez, R., Ventura, F., & Barceló, D. (2013). Occurrence and behavior of pesticides in wastewater treatment plants and their environmental impact. The Science of the Total Environment, 458, 466–76. Kolpin, D. W., Furlong, E. T., Meyer, M. T., Thurman, E. M., Zaugg, S. D., Barber, L. B., & Buxton, H. T. (2002). Pharmaceuticals, hormones, and other organic wastewater contaminants in U.S. streams, 1999–2000: a national reconnaissance. Environmental Science and Technology, 36(6), 1202–1211. Kummerer, K. (2003). Significance of antibiotics in the environment. Journal of Antimicrobial Chemotherapy, 52(1), 5–7. Kummerer, K. (2010). Pharmaceuticals in the environment. Annual Review of Environment and Resources, 35, 57–75. Le-Minh, N., Khan, S. J., Drewes, J. E., & Stuetz, R. M. (2010). Fate of antibiotics during municipal water recycling treatment processes. Water Research, 44(15), 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. Environment International, 42, 1–9. Lian, F., Song, Z., Liu, Z., Zhu, L., & Xing, B. (2013). Mechanistic understanding of tetracycline sorption on waste tire powder and its chars as affected by Cu2+ and pH. Environmental Pollution, 178, 264–270. Lillenberg, M., Herodes, K., Kipper, K., Nei, L., (2010). Plant uptake of some pharmaceuticals from fertilized soils, proceedings of the 2010 International Conference on Environmental Science and Technology, 161–165. Luo, Y., Xu, L., Rysz, M., Wang, Y., Zhang, H., & Alvarez, P. J. J. (2011). Occurrence and transport of tetracycline, sulfonamide, quinolone, and macrolide antibiotics in the Haihe River Basin, China. Environmental Science & Technology, 45, 1827–1833. Luo, Y., Guo, W., Ngo, H. H., Nghiem, L. D., Hai, F. I., Zhang, J., Liang, S., & Wang, X. C. (2014). A review on the occurrence of micropollutants in the aquatic environment and their fate and removal during wastewater treatment. Science of the Total Environment, 473–474, 619–641. Mispagel, H., & Gray, J. T. (2005). Antibiotic resistance from wastewater oxidation ponds. Water Environment Research, 77, 2996–3002. Pamreddy, A., Hidalgo, M., Havel, J., & Salvadó, V. (2013). Determination of antibiotics (tetracyclines and sulfonamides) in biosolids by pressurized liquid extraction and liquid chromatography–tandem mass spectrometry. Journal of Chromatography, A, 1298, 68–75. Saitoh, T., Shibata, K., & Hiraide, M. (2014). Rapid removal and photodegradation of tetracycline in water by surfactantassisted coagulation–sedimentation method. Journal of Environmental Chemical Engineering, 2(I3), 1852–1858.

Environ Monit Assess (2015) 187:750 Škrášková, K., Santos, L. H. M. L. M., Šatínský, D., Pena, A., Montenegro, M. C. B. S. M., Solich, P., & Nováková, L. (2013). Fast and sensitive UHPLC methods with fluorescence and tandem mass spectrometry detection for the determination of tetracycline antibiotics in surface waters. Journal of Chromatography B, 927, 201–208. Szczepanowski, R., Linke, B., Krahn, I., Gartemann, K. H., Gotzkow, T., Eichler, W., et al. (2009). Detection of 140 clinically relevant antibiotic resistance genes in the plasmid metagenome of wastewater treatment plant bacteria showing reduced susceptibility to selected antibiotics. Microbiology, 155, 2306–2319. Topal, M., & Arslan Topal, E. I. (2011). Evaluation of the Elazig municipal wastewater treatment plant with some parameters in 2010–2011 winter season. Cumhuriyet Science Journal, 32(2), 1–12. Topal, M., & Arslan Topal, E. I. (2014). Determination of the effect of municipal wastewater treatment plant effluents on the water quality of Kehli Stream. BEU Journal of Science, 3(1), 53–64. Topal, M., Uslu, G., Şahin, M., Arslan Topal, E.I., (2012). Investigation of the presence of antibiotic residues in influent of Elazig municipal wastewater treatment plant, consumer society and environment symposium. Karabük, 33–47.

Page 9 of 9 750 Topal, M., Uslu Şenel, G., Öbek, E., Arslan Topal, E.I., (2015). Bioaccumulation of tetracycline and degradation products in Lemna gibba L. exposed to secondary effluents. Desalination and Water Treatment, 1–8. doi:10.1080/19443994.2015. 1018332 Urtiaga, A. M., Pérez, G., Ibáñez, R., & Ortiz, I. (2013). Removal of pharmaceuticals from a WWTP secondary effluent by ultrafiltration/reverse osmosis followed by electrochemical oxidation of the RO concentrate. Desalination, 331, 26–34. 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 Research, 41(18), 4164–4176. Watkinson, A. J., Murby, E. J., Kolpin, D. W., & Costanzo, S. D. (2009). The occurrence of antibiotics in an urban watershed: from wastewater to drinking water. The Science of the Total Environment, 407(8), 2711–2723. Zhou, L. J., Ying, G. G., Liu, S., Zhao, J. L., Yang, B., Chen, Z. F., & Lai, H. J. (2013a). Occurrence and fate of eleven classes of antibiotics in two typical wastewater treatment plants in South China. Science of the Total Environment, 452–453, 365–376. Zhou, L. J., Ying, G. G., Liu, S., Zhang, R. Q., Lai, H. J., Chen, Z. F., & Pan, C. G. (2013b). EXcretion masses and environmental occurrence of antibiotics in typical swine and dairy cattle farms in China. Science of the Total Environment, 444, 183–195.