Contamination of surface, ground, and drinking water from pharmaceutical production* (original) (raw)

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Department of Chemistry, Umeå University, Linneausväg 6, SE‐90187 Umeå, Sweden

Department of Chemistry, Umeå University, Linneausväg 6, SE‐90187 Umeå, Sweden

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Department of Chemistry, Umeå University, Linneausväg 6, SE‐90187 Umeå, Sweden

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Department of Chemistry, Umeå University, Linneausväg 6, SE‐90187 Umeå, Sweden

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Department of Chemistry, Umeå University, Linneausväg 6, SE‐90187 Umeå, Sweden

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Department of Chemistry, Umeå University, Linneausväg 6, SE‐90187 Umeå, Sweden

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Department of Neuroscience and Physiology, Sahlgrenska Academy at the University of Gothenburg, SE‐405 30 Gothenburg, Sweden

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Received:

18 February 2009

Published:

01 December 2009

Revision received:

06 January 2010

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Jerker Fick, Hanna Söderström, Richard H. Lindberg, Chau Phan, Mats Tysklind, D. G. Joakim Larsson, Contamination of surface, ground, and drinking water from pharmaceutical production, Environmental Toxicology and Chemistry, Volume 28, Issue 12, 1 December 2009, Pages 2522–2527, https://doi.org/10.1897/09-073.1
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Abstract

Low levels of pharmaceuticals are detected in surface, ground, and drinking water worldwide. Usage and incorrect disposal have been considered the major environmental sources of these microcontaminants. Recent publications, however, suggest that wastewater from drug production can potentially be a source of much higher concentrations in certain locations. The present study investigated the environmental fate of active pharmaceutical ingredients in a major production area for the global bulk drug market. Water samples were taken from a common effluent treatment plant near Hyderabad, India, which receives process water from approximately 90 bulk drug manufacturers. Surface water was analyzed from the recipient stream and from two lakes that are not contaminated by the treatment plant. Water samples were also taken from wells in six nearby villages. The samples were analyzed for the presence of 12 pharmaceuticals with liquid chromatography‐mass spectrometry. All wells were determined to be contaminated with drugs. Ciprofloxacin, enoxacin, cetirizine, terbinafine, and citalopram were detected at more than 1 μg/L in several wells. Very high concentrations of ciprofloxacin (14 mg/L) and cetirizine (2.1 mg/L) were found in the effluent of the treatment plant, together with high concentrations of seven additional pharmaceuticals. Very high concentrations of ciprofloxacin (up to 6.5 mg/L), cetirizine (up to 1.2 mg/L), norfloxacin (up to 0.52 mg/L), and enoxacin (up to 0.16 mg/L) were also detected in the two lakes, which clearly shows that the investigated area has additional environmental sources of insufficiently treated industrial waste. Thus, insufficient wastewater management in one of the world's largest centers for bulk drug production leads to unprecedented drug contamination of surface, ground, and drinking water. This raises serious concerns regarding the development of antibiotic resistance, and it creates a major challenge for producers and regulatory agencies to improve the situation.

INTRODUCTION

The potential impacts of pharmaceutical residues in the environment have been an emerging research area during recent years. Low nanograms per liter up to a few micrograms per liter of more than 150 active pharmaceutical ingredients (APIs) have been detected in sewage effluents and surface waters, mainly as a result of excretion with urine and feces and to some extent from inappropriate disposal of unused drugs [1–6]. Pollution from the production of pharmaceuticals has not been previously considered to significantly contribute to the release of drugs into the environment; however, recent investigations of Indian and Chinese production units have shown that certain production sites are contributing to environmental pollution levels far above those previously reported [7–10]. Hyderabad is one of the world's largest hubs for the bulk drug industry, supplying Europe, the United States, and other regions with many of the most widely used generic active substances [11–12]. The Patancheru industrial area is centered approximately 25 km northwest of Hyderabad, India, and covers 120 km2; it was established in the mid‐1970s, and the industrialization has severely impacted the local environment [13]. In a recent study, Larsson et al. [9] showed extraordinarily high levels (mg/L) of several drugs in the effluent from a local wastewater treatment plant, Patancheru Enviro Tech Ltd. (PETL). This plant receives approximately 1,500 m3 of wastewater per day, primarily from approximately 90 bulk drug manufacturers, and the effluent is discharged in the Isakavagu stream feeding the Nakkavagu, Manjira, and eventually Godawari rivers [9]. In this study, we investigated whether the surface, ground, and drinking water in the region have been contaminated by antibiotics and other pharmaceuticals from bulk drug production, as this could lead to unintentional human exposure and increased risks for the development of resistance. We also investigated whether there were environmental sources for insufficiently treated industrial waste other than PETL.

MATERIALS AND METHODS

Chemicals

Ciprofloxacin, enrofloxacin, norfloxacin, ofloxacin, enoxacin, lomefloxacin, cetirizine, trimethoprim metoprolol, terbinafine, and citalopram were obtained from Sigma‐Aldrich. Enalaprilate (BP907‐F67029) was bought from European Pharmacopoeia. All drug standards were classified as high‐performance liquid chromatography (HPLC) grade (>98%). Formic acid and methanol (HPLC grade) were purchased from JT Baker, acetonitrile (HPLC grade) from Fischer Chemicals, and sulfuric acid from Merck. Purified water (resistivity, 18.2 MΩ cm) was prepared by an Elga Maxima HPLC ultrapure water system, equipped with an ultraviolet radiation source. 13C3‐trimethoprim (99%) and 13C315N‐ciprofloxacin (99%) were obtained from Cambridge Isotope Laboratories.

Fig. 1.

Map of the region, Patancheru, India, with sampling sites indicated. Patancheru Enviro Tech Ltd (PETL) = outlet from common effluent treatment plant; R1 to R5 = river water sampling sites upstream and downstream of PETL; L1 and L2 = lakes; W1 to W6 = wells located in six different villages. Arrows indicate the flow direction. For more information, see Table S1 (http://dx.doi.org/10.1897/09–073.S1).

Description of the wastewater treatment plant

Patancheru Enviro Tech Ltd., located in Patancheru near Hyderabad, India (Fig. 1), receives approximately 1,500 m3 of wastewater per day, primarily from approximately 90 bulk drug manufacturers. These industries comprise examples of the entire production chain, via synthesis of intermediates to active ingredients. The wastewater is transported on trucks to PETL, where it is collected in a buffer cistern with a retention time of approximately 2 d, thereby ensuring a more homogenous influent. Most of these cisterns are supplied with an aeration grid system and air blowers, which strips off the volatiles in the influent. After chemically assisted removal of solids, approximately 20% raw domestic sewage is added to improve biological treatment efficiency. The retention time in the aerated or oxygenated biological treatment is approximately 4 d, and levels of microorganisms are maintained by recirculation of sludge. After settling and centrifugation, the clarified water is discharged into the Isakavagu stream. The content of organic material measured as biological oxygen demand and chemical oxygen demand is reported to be reduced from typically 1,300 and 6,000 mg/L, respectively, in the mixed influent to 270 and 1,400 mg/L in the treated effluent [9]. Total dissolved solids and total suspended solids are reduced from approximately 9,000 and 500 to 5,000 and 300 mg/L, respectively [9].

Sampling and sample preparation

Grab samples were taken approximately 0.1 m below the surface in duplicate at the same sampling occasion. The samples were filtered through a 0.45‐μm membrane filter (Millipore) and adjusted to pH 3 using sulfuric acid. Solid‐phase extraction (SPE) columns (Oasis HLB, 200 mg, Waters) were preconditioned and equilibrated with 5.0 ml of methanol and 5.0 ml of deionized water. We applied 100 ml of each water sample to the SPE columns at a flow rate of 5 ml min−1 within 12 h of sampling and stored them at 4°C prior to and during shipment. Detailed information regarding sampling points and sampling dates is presented in the Table S1 (http://dx.doi.org/10.1897/09–073.S1). Prior to elution, SPE columns were reconditioned with 2 ml of deionized water. As surrogate standards, 500 ng of 13C315N and 13C3, labeled ciprofloxacin and trimethoprim, respectively, were used. These were applied to the SPE columns before elution with 5 ml of methanol. All samples were eluted within one month after sampling. Eluates were evaporated under a gentle air stream to a volume of 20 μL before being reconstituted in 1 ml of the initial mobile phase (95% water, 5% acetonitrile, and 0.1% [v/v] formic acid).

Liquid chromatography‐mass spectrometry

Aliquots (10 μl) of sample extracts and calibration solutions were injected into a YMC Hydrosphere C18 analytical column, 150 × 4.6 mm internal diameter, 5‐μm particle size, following a 10 × 4 mm internal diameter, 5‐μm particle size, YMC Hydrosphere C18 guard column using an AS 3000 auto injector (Thermo Finnigan). Water and acetonitrile, both containing 0.1% (v/v) formic acid, were used as mobile phase. Analytes were chromatographically separated with a gradient from 95% water and 5% acetonitrile to 40% water and 60% acetonitrile for 10 min and held for 1.5 min before being brought back to the initial condition. Analysis was done at a flow rate of 0.8 ml min−1 generated by a P4000 HPLC pump (Spectra system, Thermo Finnigan) at 25°C. An LCQ Duo ion‐trap mass spectrometer (Thermo Finnigan) was used, together with an electrospray ion source in positive ion mode. The source voltage was maintained at a constant 6.0 kV, and the heated capillary temperature was set to 250°C. The tandem mass spectrometry parameters were semiautomatically optimized for the analytes using the LCQ Duo internal software; the collision energy, for producing daughter ions, was optimized manually. Liquid chromatography coupled to tandem mass spectrometry with electrospray ionization parameters are presented in Table 1.

Surrogate standards, identification, and quantification

In the present study, 13C‐labeled standards were used to improve accuracy in the quantification of the analytes. Identification of each analyte was based on its chromatographic retention time, and selectivity was ensured by using a single appropriate transition in tandem mass spectrometry mode. Labeled ciprofloxacin was used for the determination of all fluoroquinolones, and labeled trimethoprim was used for the determination of trimethoprim, cetirizine, citalopram, enalapril, metoprolol, and terbinafine. The analytes were quantified by selected reaction monitoring, with the most abundant daughter ion recorded in each case. The internal standard calibration method was used for all aqueous samples; analyte to surrogate standard peak area ratios were calculated. To cover the very large concentration range, three separate five‐point calibration curves were used for quantification, generated from standards prepared in water to acetonitrile (95:5 v/v) and 0.1% (v/v) formic acid, with concentrations ranging from 1 ng ml−1 to 25 mg ml−1 of the active substances.

Quality assurance

Stock solutions were prepared every three months and stored at 4°C. The recoveries of the analytes in the fortified laboratory samples were evaluated with 200 ml of tap water, river water, and effluent, fortified with 1 μg of each substance. Spiked samples were loaded into Oasis HLB columns, previously conditioned by 5 ml of methanol, 5 ml of a methanol to water ratio (50:50), and 5 ml of water adjusted to pH 3, at flow rates of approximately 5 ml min−1. Solid‐phase extraction cartridges included in the storage study were transferred to dark glass bottles and maintained at 4°C before analysis. Recoveries were determined from the analyte to surrogate standard peak area ratios for the spiked samples in a matching standard calibration curve. Field blanks were also analyzed via the same analytical protocol. Field blanks were transported and stored together with the other samples, both during the analysis in Sweden and during the sampling in India. Injection of the mobile phase was performed regularly during the analysis to detect potential carryover effects.

Table 1.

Included pharmaceuticals and LC‐ESI‐MS/MSa parametersbc

a Liquid chromatography coupled to tandem mass spectrometry with electrospray ionization.

b The source voltage was maintained at a constant 6.0 kV and the heated capillary temperature was set to 250°C.

c CE = collision energy; a.u. = arbitrary units; LOQ = limit of quantification.

d Slashes separate different daughter ions used.

Table 1.

Included pharmaceuticals and LC‐ESI‐MS/MSa parametersbc

a Liquid chromatography coupled to tandem mass spectrometry with electrospray ionization.

b The source voltage was maintained at a constant 6.0 kV and the heated capillary temperature was set to 250°C.

c CE = collision energy; a.u. = arbitrary units; LOQ = limit of quantification.

d Slashes separate different daughter ions used.

RESULTS

Analytes were considered to have been positively identified when the following criteria were met: Selected reaction monitoring fragmentation patterns were consistent with those shown in Table 1, and analyte retention times were within 2% relative standard deviation of the average times for the corresponding analyte in calibration standards. The calibration curves of the monitored antibiotics showed good linearity, with a minimum regression coefficient of 0.97, with one exception: The regression coefficient for the calibration curve (1,000–10,000 ng ml−1) for enaprilate was 0.94. The recoveries were above 85% for all pharmaceuticals after one month of storage at 4°C in river water (Table S2; http://dx.doi.org/10.1897/09–073.S1). The limits of quantitation were determined by using the second point in the lowest calibration curve for each analyte; peak areas had to have a signal to noise ratio of 10 or greater (Table 1). A chromatogram of one of the river samples is shown in Figure S1 (http://dx.doi.org/10.1897/09–073.S1). The analyte levels in the blank samples were not detectable (less than the limit of quantification), and no carryover effects were noticed.

Table 2.

Levels of pharmaceuticals in potential sources, average of two samples from each site (ng/L)a‐c

a PETL = Patancheru Enviro Tech Ltd., undiluted effluent from the treatment plant.

c ND = not detected.

Table 2.

Levels of pharmaceuticals in potential sources, average of two samples from each site (ng/L)a‐c

a PETL = Patancheru Enviro Tech Ltd., undiluted effluent from the treatment plant.

c ND = not detected.

The effluent from PETL contained 14 mg/L of ciprofloxacin and 2.1 mg/L of cetirizine and microgram per liter levels of several additional pharmaceuticals (Table 2), which is in accordance with results presented earlier [9]. Samples taken upstream from PETL (R1 and R2) contained nanogram per liter to low microgram per liter levels of the analyzed pharmaceuticals (Table 3). Downstream samples showed a declining gradient, with ciprofloxacin (2,500 to 10 μg/L) apparently decreasing in concentration at the furthest sampling point more than, for example, cetirizine (530 to 97 μg/L), and citalopram (76 to 0.5 μg/L; Table 3).

Lakes 1 and 2 are located upstream of PETL on a tributary stream (Fig. 1). Ciprofloxacin was not detected in L1, but elevated levels of other pharmaceuticals were detected, including norfloxacin (520 μg/L) and terbinafine (15 μg/L; Table 2). Lake 1 was partially dried out during the sampling period; thus, two samples (L1a and L1b) were taken from the largest water body present and one sample (L1c) was taken from a smaller, adjacent water body. Lake 2 is situated immediately north of a pharmaceutical production area. Ciprofloxacin levels between 2.5 and 6.5 mg/L were measured, with the highest levels recorded closest to the industrial area (L2c; Table 2). Cetirizine was found in levels between 1.2 and 0.34 mg/L (Table 2).

Table 3.

Levels of pharmaceuticals in the Isakavagu‐Nakkavagu rivers, India, based on an average of two samples from each site (ng/L)ab

a R1 is 3 km upstream from Patancheru Enviro Tech Ltd. (PETL); R2 is 1 km upstream from PETL; R3 is 150 m downstream from PETL; R4 is 4 km downstream from PETL; and R5 is 30 km downstream from PETL. For the locations of sampling points R1 to R5 and sampling dates, see Figure 1 and Table S1 (http://dx.doi.org/10.1897/09–073.S1).

b ND = not detected.

Table 3.

Levels of pharmaceuticals in the Isakavagu‐Nakkavagu rivers, India, based on an average of two samples from each site (ng/L)ab

a R1 is 3 km upstream from Patancheru Enviro Tech Ltd. (PETL); R2 is 1 km upstream from PETL; R3 is 150 m downstream from PETL; R4 is 4 km downstream from PETL; and R5 is 30 km downstream from PETL. For the locations of sampling points R1 to R5 and sampling dates, see Figure 1 and Table S1 (http://dx.doi.org/10.1897/09–073.S1).

b ND = not detected.

Wells 1 to 3 are located close to the investigated lakes, whereas W4 to W6 are close (400–600 m) to the Nakkavagu River, which receives water from the two lakes and from the Isakavagu stream, where the PETL effluent is discharged (Fig. 1). All wells contained high nanogram or low microgram per liter levels of several of the investigated pharmaceuticals (Table 4). According to local people the well with the highest levels (W2, up to 14 and 28 μg/L of ciprofloxacin and cetirizine, respectively) and W5 are no longer used for human consumption, whereas W1, W3, W4, and W6 are currently used as drinking‐water sources as other supplies are often not sufficient.

DISCUSSION

The present study clearly shows that pharmaceutical production severely contaminates surface, ground, and drinking water in the investigated region and that the previously demonstrated release of pharmaceutical residues from PETL is still occurring. The contamination of the lakes with milligrams per liter of drugs was a 100,000 to 1 million times higher than reported levels of fluoroquinolones in surface water in the United States and China contaminated by sewage effluents [2,14]. This shows that effluent from PETL is not the only environmental source for insufficiently treated industrial effluents containing high levels of APIs. Indeed, the Andhra Pradesh Pollution Control Board acknowledges that the unauthorized dumping of industrial waste has been a significant problem in the region [15; www.toxicslink.org/docs/SCMC_Visit_AP.doc], but dumped waste has not been analyzed for drug residues. In one of the lakes (L2), ciprofloxacin and cetirizine levels even exceeded human therapeutic blood plasma concentrations of 2.5 mg/L and 20 μg/L for ciprofloxacin and cetirizine, respectively [16]. The analysis of the well water indicates that the analyzed drugs can contaminate groundwater over large areas, which presents a direct route for human exposure. To the best of our knowledge, these surface‐ and well‐water levels are the highest reported to date. It may not seem economically viable to discharge milligrams per liter of pharmaceuticals; however, calculations show that the cost of purchasing bulk pharmaceuticals is often only a small fraction of the price of the final pharmaceutical products [9,11]. For ciprofloxacin, the cost of purchasing the API accounts for less than 1.5% of the sales price of the final product in Sweden [11]. This should be compared with the investment and operating costs for producing a clean effluent.

Table 4.

Levels of pharmaceuticals in wells, average of two samples from each site (ng/L)ab

b ND = not detected.

c Single sample, duplicate lost.

Table 4.

Levels of pharmaceuticals in wells, average of two samples from each site (ng/L)ab

b ND = not detected.

c Single sample, duplicate lost.

The most obvious risk associated with the findings in the present study is that the high levels of broad‐spectrum antibiotics could induce the development of antibiotic‐resistant microorganisms. The increasing occurrence of multiresistant pathogens is a serious global threat to human health and is promoted by the heavy use of antibiotics in human and veterinary medicine. Environmental bacteria constitute a potential recruitment pool for resistance factors that, through horizontal gene transfer, may end up in human pathogens [17,18], rendering current treatment regimens useless. It has been shown previously that industrial wastewater containing antibiotics can select for resistant strains of bacteria in the environment [19]. Li et al. [20] recently showed that the proportions of multiresistant bacterial strains sampled inside and downstream from a manufacturing site of penicillin in China was higher than in upstream isolates. Resistance to β‐lactam antibiotics, such as penicillin, was very frequent, and generally the strains showed a higher minimal inhibitory concentration for the β‐lactams than for other classes of antibiotics. Many strains were highly multiresistant, suggesting that exposure to one antibiotic class can select for resistance to other classes of antibiotics as well [20]. Even much lower environmental levels of antibiotics, originating from human usage, have recently raised concerns for resistance development [17,21], but the risks are apparently much higher at certain manufacturing sites.

The PETL applies an activated sludge treatment in which, at the end of the process, some surviving bacteria are recycled to select those that are good at metabolizing the influent. This is normally a desirable procedure for improving the treatment efficiency at a sewage treatment plant; however, with high levels of antibiotics in the influent, it also means an active selection for resistant bacteria. In addition, approximately 20% of raw human feces, inevitably containing pathogens, are added daily to maintain biological activity. Such a close contact among pathogens, resistant bacteria, and antibiotics can facilitate a transfer of resistance to pathogens. The addition of raw sewage may have contributed some pharmaceutical residues to the effluent but is not sufficient to explain the high levels reported here and in our earlier study [9]. The present study also shows that very large volumes of surface and groundwater are contaminated by fluoroquinolones at levels high enough (<5–10 μg/L) to promote horizontal transfer of resistance genes [22]. The use of contaminated groundwater as a drinking‐water source could therefore act as a direct route for resistant bacteria to reach humans. Thus, it is important to perform studies on the environmental microflora in the Patancheru area to assess the full impact of the fluoroquinolone contamination.

Direct exposure to pharmaceuticals at levels normally found in drinking water (up to 100 ng/L) is generally not considered to pose human health risks [23–25]. However, the present study shows that drinking water in areas with high levels of industrial waste may be contaminated to considerably higher levels. Although the estimated daily exposure would still be far below normal therapeutic doses for all analyzed drugs, the risk in such areas, particularly regarding exposure during pregnancy and childhood [26], should be investigated further. It is also evident that the levels of fluoroquinolones measured will have ecotoxicological effects, particularly on microbial ecosystems [9,27,28]. As bacteria play important roles in the cycling of energy and nutrients, effects on microbial ecology may indirectly have unanticipated consequences for other parts of the ecosystems. However, the effluent from PETL also contains substances that directly affect vertebrates; for example, the growth of frog tadpoles exposed to the effluent diluted 1:500 was strongly impaired [29].

Several factors influence the fate and mobility of pharmaceuticals; thus, the classical Karickhoff's formula correlating the organic carbon based sorption coefficient with the octanol‐water coefficient is difficult to apply directly on many drugs [30,31]. Fluoroquinolones bind strongly, despite low octanol‐water coefficient values, to sludge in regular sewage treatment plants [3], probably through metal complexes interactions [32]. Given the high concentrations of ciprofloxacin in the effluent, one could speculate that levels in the sludge from PETL would have been considerably higher, although this was not analyzed in the present study. Some pharmaceuticals are known to pass through the floor of rivers and lakes and mix with groundwater [33,34], and Massmann et al. recently showed that pharmaceuticals can persist in aquatic environments for decades [35]. As exposure to sunlight is considered the most important factor by which fluoroquinolones are degraded in nature [36,37], it is likely that fluoroquinolones are relatively stable once they reach the groundwater. It is clear that the groundwater in the region is contaminated by the effluent from PETL and water from L1 and L2, but it is difficult to elucidate the exact fate of drugs spreading from these point sources. Factors complicating the interpretation are that the geochemistry in the region not is fully characterized and that some of the samples were taken on different dates. Elevated levels of ciprofloxacin and enoxacin in W2, W3, and L2, however, suggest a common contamination source. Although ciprofloxacin was more abundant than cetirizine both in the effluent from PETL and in L1, all six investigated wells had higher levels of cetirizine than ciprofloxacin. This could indeed be expected based on the high water solubility of cetirizine and the known ability of fluoroquinolones, including ciprofloxacin, to bind to sludge and sediments [3,38]. Indeed ofloxacin, another fluoroquinolone, adsorbs strongly to different types of soils and is not released easily [29]. The detection of ciprofloxacin (up to 14 μg L−1), enoxacin (up to 1.9 μg/L), and ofloxacin (up to 0.48 μg/L) in the well water samples nevertheless demonstrate a significant transport of fluoroquinolones with the groundwater (Table 4).

The severe contamination of surface, ground‐, and drinking water with pharmaceuticals described here undoubtedly affects the local ecosystems. However, in our view, the most urgent aspect of the environmental drug contamination is that high levels of broad‐spectrum antibiotics are likely to promote the development of highly antibiotic‐resistant microorganisms and possibly horizontal transfer of resistance factors to human pathogens. High levels of fluoroquinolones, activated sludge technology, the addition of human feces to the treatment process at PETL, and the very large contaminated volumes of surface and groundwater from different sources add to this risk. Bulk drugs produced in the Patancheru area are exported to other countries, including the major markets in Europe and the United States [11]. Larsson and Fick [12] recently identified the sources of the APIs in all 242 medicinal products on the Swedish market containing any of nine preselected APIs. In 74 of the products (31%), the API originated from Indian producers sending their wastewater to PETL. This implies an international responsibility to improve the current environmental situation in Patancheru, especially considering that the fight against antibiotic resistance is a worldwide challenge.

SUPPORTING INFORMATION

Table S1. Detailed sample information.

Table S2. Recoveries in different matrices and after storage.

Fig. S1. Chromatogram of river sample R1.

All found at DOI: 10.1897/09–073.S1 (92 KB PDF).

Acknowledgements

We thank S. Strindberg, L. Svensson, and the Gamana Organization for input and assistance with sampling and the Foundation for Strategic Environmental Research, Swedish Research Council for Environment, Agricultural Sciences and Spatial Planning, Swedish Research Council, Swedish International Development Cooperation Agency, and Adlerbertska Foundation for financial support.

REFERENCES

1

Heberer

T

.

2002

.

Occurrence, fate, and removal of pharmaceutical residues in the aquatic environment: A review of recent research data

.

Toxicol Lett

131

:

5

–

17

.

2

Kolpin

DW

,

Furlong

ET

,

Meyer

MT

,

Thurman

EM

,

Zaugg

SD

,

Barber

LB

,

Buxton

HT

.

2002

.

Pharmaceuticals, hormones, and other organic wastewater contaminants in U.S. streams, 1999–2000: A national reconnaissance

.

Environ Sci Technol

36

:

1202

–

1211

.

3

Lindberg

RH

,

Wennberg

P

,

Johansson

MI

,

Tysklind

M

,

Andersson

BAV

.

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

.

4

Fent

K

,

Weston

AA

,

Caminada

D

.

2006

.

Ecotoxicology of human pharmaceuticals

.

Aquat Toxicol

76

:

122

–

159

.

5

Williams

RT

,

Cook

JC

.

2007

.

Exposure to pharmaceuticals present in the environment

.

Drug Inf J

41

:

133

–

141

.

6

Nikolaou

A

,

Meric

S

,

Fatta

D

.

2007

.

Occurrence patterns of pharmaceuticals in water and wastewater environments

.

Anal Bioanal Chem

387

:

1225

–

1234

.

7

Bisarya

SC

,

Patil

DM

.

1993

.

Determination of salicylic‐acid and phenol (ppm level) in effluent from aspirin plant

.

Research and Industry

38

:

170

–

172

.

8

Cui

CW

,

Ji

SL

,

Ren

HY

.

2006

.

Determination of steroid estrogens in wastewater treatment plant of a controceptives producing factory

.

Environ Monit Assess

121

:

409

–

419

.

9

Larsson

DGJ

,

de Pedro

C

,

Paxeus

N

.

2007

.

Effluent from drug manufactures contains extremely high levels of pharmaceuticals

.

J Hazard Mater

148

:

751

–

755

.

10

Li

D

,

Yang

M

,

Hu

J

,

Ren

L

,

Zhang

Y

,

Li

K

.

2008

.

Determination and fate of oxytetracycline and related compounds in oxytetracycline production wastewater and the receiving river

.

Environ Toxicol Chem

27

:

80

–

86

.

11

Larsson

DGJ

.

2008

. Drug production facilities: An overlooked discharge source for pharmaceuticals to the environment. In

Kümmerer

K

, ed,

Pharmaceuticals in the Environment. Sources, Fate, Effects and Risks

.

Springer

,

New York, NY, USA

.

12

Larsson

DGJ

,

Fick

J

.

2009

.

Transparency throughout the production chain: A way to reduce pollution from the manufacturing of pharmaceuticals?

.

Regul Toxicol Pharm

53

:

161

–

163

.

13

Rao

WSG

,

Dhar

RL

,

Subrahmanyam

K

.

2001

.

Assessment of contaminant migration in groundwater from an industrial development area, Medak district, Andhra Pradesh, India

.

Water Air Soil Pollut

128

:

369

–

389

.

14

Xiao

Y

,

Chang

H

,

Jia

A

,

Hu

JY

.

2008

.

Trace analysis of quinolone and fluoroquinolone antibiotics from wastewaters by liquid chromatography‐electrospray tandem mass spectrometry

.

J Chromatogr A

1214

:

100

–

108

.

15

Boralkar

DB

,

Alvares

C

,

Devotta

S

,

Sharma

PN

,

Thyagarajan

G

.

2004

. Report of visit to Hyderabad (A.P.) October 19–20. Supreme Court Monitoring Committee on Hazardous Wastes. Supreme Court of India, New Delhi.

16

Schulz

M

,

Schmoldt

A

.

2003

.

Therapeutic and toxic blood concentrations of more than 800 drugs and other xenobiotics

.

Pharmazie

58

:

447

–

474

.

17

Martinez

JL

.

2008

.

Antibiotics and antibiotic resistance genes in natural environments

.

Science

321

:

365

–

367

.

18

Dantas

G

,

Sommer

MOA

,

Oluwasegun

RD

,

Church

GM

.

2008

.

Bacteria subsisting on antibiotics

.

Science

320

:

100

–

103

.

19

Guardabassi

L

,

Petersen

A

,

Olsen

JE

,

Dalsgaard

A

.

1998

.

Antibiotic resistance in Acinetobacter spp. isolated from sewers receiving waste effluent from a hospital and a pharmaceutical plant

.

Appl Environ Microbiol

64

:

3499

–

3502

.

20

Li

D

,

Yang

M

,

Hu

J

,

Zhang

J

,

Liu

R

,

Gu

X

,

Zhang

Y

,

Wang

Z

.

2009

.

Antibiotic‐resistance profile in environmental bacteria isolated from penicillin production wastewater treatment plant and the receiving river

.

Environ Microbiol

11

:

1506

–

1517

.

21

Cattoir

V

,

Poirel

L

,

Aubert

C

,

Soussy

CJ

,

Nordmann

P

.

2008

.

Unexpected occurrence of plasmid‐mediated quinolone resistance determinants in environmental Aeromonas spp

.

Emerg Infect Dis

14

:

231

–

237

.

22

Beaber

JW

,

Hochhut

B

,

Waldor

MK

.

2004

.

SOS response promotes horizontal dissemination of antibiotic resistance genes

.

Nature

427

:

72

–

74

.

23

Webb

S

,

Ternes

T

,

Gibert

M

,

Olejniczak

K

.

2003

.

Indirect human exposure to pharmaceuticals via drinking water

.

Toxicol Lett

142

:

157

–

167

.

24

Stackelberg

PE

,

Furlong

ET

,

Meyer

MT

,

Zaugg

SD

,

Henderson

AK

,

Reissman

DB

.

2004

.

Persistence of pharmaceutical compounds and other organic wastewater contaminants in a conventional drinking‐water treatment plant

.

Sci Total Environ

329

:

99

–

113

.

25

Johnson

AC

,

Jurgens

MD

,

Williams

RJ

,

Kümmerer

K

,

Korten‐kamp

A

,

Sumpter

JP

.

2008

.

Do cytotoxic chemotherapy drugs discharged into rivers pose a risk to the environment and human health? An overview and UK case study

.

J Hydrol

348

:

167

–

175

.

26

Collier

AC

.

2007

.

Pharmaceutical contaminants in potable water: Potential concerns for pregnant women and children

.

Ecohealth

4

:

164

–

171

.

27

Backhaus

T

,

Scholze

M

,

Grimme

LH

.

2000

.

The single substance and mixture toxicity of quinolones to the bioluminescent bacteriumVibrio fischeri

.

Aquat Toxicol

49

:

49

–

61

.

28

Robinson

AA

,

Belden

JB

,

Lydy

MJ

.

2005

.

Toxicity of fluoroquinolone antibiotics to aquatic organisms

.

Environ Toxicol Chem

24

:

423

–

430

.

29

Carlsson

G

,

Örn

S

,

Larsson

DGJ

.

2009

.

Effluent from bulk drug production is toxic to aquatic vertebrates

.

Environ Toxicol Chem

28

:

2656

–

2662

.

30

Drillia

P

,

Stamatelatou

K

,

Lyberatos

G

.

2005

.

Fate and mobility of pharmaceuticals in solid matrices

.

Chemosphere

60

:

1034

–

1044

.

31

Yamamoto

H

,

Nakamura

Y

,

Moriguchi

S

,

Nakamura

Y

,

Honda

Y

,

Tamura

I

,

Hirata

Y

,

Hayashi

A

,

Sekizawa

J

.

2009

.

Persistence and partitioning of eight selected pharmaceuticals in the aquatic environment: Laboratory photolysis, biodegradation, and sorption experiments

.

Water Res

43

:

351

–

362

.

32

Aristilde

L

,

Sposito

G

.

2008

.

Molecular modeling of metal complexation by a fluoroquinolone antibiotic

.

Environ Toxicol Chem

27

:

2304

–

2310

.

33

Reddersen

K

,

Heberer

T

,

Dunnbier

U

.

2002

.

Identification and significance of phenazone drugs and their metabolites in ground and drinking water

.

Chemosphere

49

:

539

–

544

.

34

Heberer

T

.

2002

.

Tracking persistent pharmaceutical residues from municipal sewage to drinking water

.

J Hydrol

266

:

175

–

189

.

35

Massmann

G

,

Sultenfuss

J

,

Dunnbier

U

,

Knappe

A

,

Taute

T

,

Pekdeger

A

.

2008

.

Investigation of groundwater a residence times during bank filtration in Berlin: Multi‐tracer approach

.

Hydrol Process

22

:

788

–

801

.

36

Burhenne

J

,

Ludwig

M

,

Nikoloudis

P

,

Spiteller

M

.

1997

.

Photolytic degradation of fluoroquinolone carboxylic acids in aqueous solution. I. Primary photoproducts and half‐lives

.

Environ Sci Pollut Res

4

:

10

–

15

.

37

Boreen

AL

,

Arnold

WA

,

McNeill

K

.

2003

.

Photodegradation of pharmaceuticals in the aquatic environment: A review

.

Aquat Sci

65

:

320

–

341

.

38

Pico

Y

,

Andreu

V

.

2007

.

Fluoroquinolones in soil: Risks and challenges

.

Anal Bioanal Chem

387

:

1287

–

1299

.

Author notes

Published on the Web 5/18/2009.

Copyright © 2009 SETAC

This article is published and distributed under the terms of the Oxford University Press, Standard Journals Publication Model (https://academic.oup.com/pages/standard-publication-reuse-rights)

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