(1) Eco Toxicology of Human Pharmaceuticals (FENT - 2006)

Aquatic Toxicology 76 (2006) 122–159

Review

Ecotoxicology of human pharmaceuticals

Karl Fenta,b,∗, Anna A. Westona,c, Daniel Caminadaa,d

a University of Applied Sciences Basel, Institute of Environmental Technology, St. Jakobs-Strasse 84, CH-4132 Muttenz, Switzerlandb Swiss Federal Institute of Technology (ETH), Department of Environmental Sciences, CH-8092 Zurich, Switzerland

c Springborn Smithers Laboratories Europe AG, Seestrasse 21, CH-9326 Horn, Switzerlandd University of Zurich, Institute of Plant Biology, Limnology, Seestrasse 187, CH-8802 Kilchberg, Switzerland

Received 21 February 2005; received in revised form 1 September 2005; accepted 1 September 2005

Abstract

Low levels of human medicines (pharmaceuticals) have been detected in many countries in sewage treatment plant (STP)effluents, surface waters, seawaters, groundwater and some drinking waters. For some pharmaceuticals effects on aquatic organ-isms have been investigated in acute toxicity assays. The chronic toxicity and potential subtle effects are only marginally known,however. Here, we critically review the current knowledge about human pharmaceuticals in the environment and address severalkey questions. What kind of pharmaceuticals and what concentrations occur in the aquatic environment? What is the fate insurface water and in STP? What are the modes of action of these compounds in humans and are there similar targets in loweranimals? What acute and chronic ecotoxicological effects may be elicited by pharmaceuticals and by mixtures? What are theeffect concentrations and how do they relate to environmental levels? Our review shows that only very little is known about

humanchronic

her thanentra-asuredtargetedl effects.

23125

long-term effects of pharmaceuticals to aquatic organisms, in particular with respect to biological targets. For mostmedicines analyzed, acute effects to aquatic organisms are unlikely, except for spills. For investigated pharmaceuticalslowest observed effect concentrations (LOEC) in standard laboratory organisms are about two orders of magnitude higmaximal concentrations in STP effluents. For diclofenac, the LOEC for fish toxicity was in the range of wastewater conctions, whereas the LOEC of propranolol and fluoxetine for zooplankton and benthic organisms were near to maximal meSTP effluent concentrations. In surface water, concentrations are lower and so are the environmental risks. However,ecotoxicological studies are lacking almost entirely and such investigations are needed focusing on subtle environmentaThis will allow better and comprehensive risk assessments of pharmaceuticals in the future.© 2005 Elsevier B.V. All rights reserved.

Keywords: Pharmaceuticals; Ecotoxicological effects; Environmental toxicity; Chronic effects; Environmental risk assessment

Contents

1. Introduction. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12. Sources. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

∗ Corresponding author.E-mail address: [email protected] (K. Fent).

0166-445X/$ – see front matter © 2005 Elsevier B.V. All rights reserved.doi:10.1016/j.aquatox.2005.09.009

K. Fent et al. / Aquatic Toxicology 76 (2006) 122–159 123

3. Fate in the environment. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1274. Environmental concentrations. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 130

4.1. Analgesics and antiinflammatory drugs. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1324.2. Beta-blockers. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1334.3. Blood lipid lowering agents. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1334.4. Neuroactive compounds (antiepileptics, antidepressants). . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1334.5. Antineoplastics and antitumor agents. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1334.6. Various other compounds. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1344.7. Steroidal hormones. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 134

5. Modes of actions in humans and mammals and occurrence of target biomolecules in lowervertebrates and invertebrates. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1355.1. Analgesics and non-steroidal antiinflammatory drugs (NSAID). . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1355.2. Beta-blockers. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1365.3. Blood lipid lowering agents. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1375.4. Neuroactive compounds (antiepileptics, antidepressants). . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1385.5. Cytostatics compounds and cancer therapeutics. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1385.6. Various compounds. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 138

6. Ecotoxicological effects. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1396.1. Acute effects. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 139

6.1.1. Analgesics and non-steroidal antiinflammatory drugs (NSAID). . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1406.1.2. Beta-blockers. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1406.1.3. Blood lipid lowering agents. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1416.1.4. Neuroactive compounds (antiepileptics, antidepressants). . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1416.1.5. Cytostatic compounds and cancer therapeutics. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 142

6.2. Chronic effects. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1426.2.1. Analgesics and non-steroidal antiinflammatory drugs. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1446.2.2. Beta-blockers. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1446.2.3. Blood lipid lowering agents. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1446.2.4. Neuroactive compounds. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 144

6.3. In vitro studies. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1466.4. Toxicity of pharmaceutical mixtures and community effects. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 146

7. Comparison of environmental concentrations and ecotoxicological effects concentrations. . . . . . . . . . . . . . . . . . . . . . 147148

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8. Discussion. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .9. Conclusions and future directions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

Acknowledgements. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .References. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

. Introduction

It came as a surprise when an unusually high deathate among three species of vulture in India and Pak-stan was reported in 2004 to be caused by diclofenac,

widely used analgesic and antiinflammatory drugOaks et al., 2004). The oriental white-backed vul-ure (Gyps bengalensis) is one of the most commonaptors in the Indian subcontinent and a populationecline of >95% makes this species as being criticallyndangered. Whereas a population decline has started

n the 1990s, recent catastrophic declines also involve

Gyps indicus andGyps tenuirostris across the Indiansubcontinent (Prakash et al., 2003; Risebrough, 200).High adult and subadult mortality and resulting poplation loss is associated with renal failure and viscegout, the accumulation of uric acid throughout the bocavity following kidney malfunction. A direct correlation between residues of diclofenac and renal failwas reported both by experimental oral exposurethrough feeding vultures diclofenac-treated livestoHence, the residues of diclofenac were made respsible for the population decline (Oaks et al., 2004).This drug has recently come into widespread use

124 K. Fent et al. / Aquatic Toxicology 76 (2006) 122–159

these countries as a veterinary medicine, but is alsowidely used as in human medicine since the 1970s.Vultures are natural scavengers feeding on carrion ofwildlife and domestic livestock and cattle. The threevulture species continue to decline in Pakistan, India,Bangladesh and southern Nepal. Apart from this severecase, never having been anticipated, potential ecotox-icological effects of drug residues in the environmenton wildlife are largely unknown.

Pharmaceuticals are a class of emerging environ-mental contaminants that are extensively and increas-ingly being used in human and veterinary medicine.These chemicals are designed to have a specific modeof action, and many of them for some persistence in thebody. These features among others make pharmaceu-ticals to be evaluated for potential effects on aquaticflora and fauna. The current investigations are mainlydriven by advances in environmental residue analy-sis, particularly after the establishment of chemicalanalysis methods able to determine more polar com-pounds such as liquid chromatography–tandem massspectrometry, which allows the identification of tracequantities of polar organic pollutants without derivati-zation (Ternes et al., 1998, 2001; Kolpin et al., 2002;Kummerer, 2004). Accordingly, many environmentalanalyses have been performed in various countries,which are summarized by various reports (e.g.Halling-Sorensen et al., 1998; Daughton and Ternes, 1999;Kummerer, 2004). These monitoring studies demon-strate that drug residues in treated wastewater and sur-f

gi-c strialo one msa osedv darda um-b mayn tiono haz-a ko ctst e oft o thee riska fewy uide-

lines on how pharmaceuticals should be assessed forpossible unwanted effects on the environment. The firstrequirement for ecotoxicity testing as a prerequisitefor registration of pharmaceuticals was established in1995 according to the European Union (EU) Directive92/18 EEC and the corresponding “Note for Guid-ance” (EMEA, 1998) for veterinary pharmaceuticals.The European Commission released a draft guide-line (Directive 2001/83/EC) specifying that an autho-rization for a medicinal product for human use mustbe accompanied by an environmental risk assessment(EMEA, 2005). The U.S. Food and Drug Administra-tion (FDA) published a guidance for the assessments ofhuman drugs; according to this, applicants in the U.S.A.are required to provide an environmental assessmentreport when the expected introduction concentrationof the active ingredient of the pharmaceutical in theaquatic environment is≥1�g/L (FDA-CDER, 1998),which corresponds to about 40 t as a trigger level.In contrast, environmental assessments of veterinarypharmaceuticals is required by the U.S. FDA since1980 (Boxall et al., 2003).

The objective of our paper is to compile and crit-ically review the present knowledge about the envi-ronmental occurrence and fate of human pharmaceu-ticals in the aquatic environment, to discuss poten-tial mechanisms of action based on knowledge frommammalian studies, and to describe the acute andchronic ecotoxicological effects on aquatic organisms.We also identify major gaps in the current knowl-e te onp e ofw ebyf ndsb on-s oodl neu-r heirm theirc thee ntale ne h eta int i-n lle ledi

ace water are very widespread.In contrast, only little is known about ecotoxicolo

al effects of pharmaceuticals on aquatic and terrerganisms and wildlife, and a comprehensive reviewcotoxicological effects is lacking. Aquatic organisre particularly important targets, as they are expia wastewater residues over their whole life. Stancute ecotoxicity data have been reported for a ner of pharmaceuticals, however, such data aloneot be suitable for specifically addressing the quesf environmental effects, and subsequently in therd and risk assessment (Fent, 2003). The current lacf knowledge holds in particular for chronic effe

hat have only very rarely been investigated. In spithe sizeable amounts of human drugs released tnvironment, concise regulations for ecologicalssessment are largely missing. Only in the lastears, regulatory agencies have issued detailed g

dge and future research needs. We concentraharmaceuticals used in human medicine, somhich are also applied in veterinary medicine, ther

ocusing on environmentally important compouelonging to different drug categories, namely nteroidal antiinflammatory drugs, beta-blockers, blipid lowering agents, cancer therapeutics andoactive compounds. These classes differ for todes of actions and were chosen because of

onsumption volumes, toxicity and persistence innvironment. We will not address the environmeffects of antibiotics and biocides (Halling-Sorenset al., 1998; Daughton and Ternes, 1999; Hirscl., 1999), hormones (used in contraceptives and

herapy) (Damstra et al., 2002) and special veterary pharmaceuticals (Montforts et al., 1999; Boxat al., 2003) as the cited reports provide detai

nformation.


The current knowledge indicates that residues ofpharmaceuticals at trace quantities are widespreadin aquatic systems. Pharmaceuticals in the environ-ment are suggested to pose only a low risk for acutetoxicity. For chronic effects, the situation may bedifferent, but there is a considerable lack of infor-mation. Investigation of multigenerational life-cycleeffects or at various life stages is lacking, althoughmany aquatic organisms are exposed for their entirelife. There is a need to focus on long-term expo-sure assessment regarding specific modes of actionof pharmaceuticals to better judge the implicationsof pharmaceutical residues in aquatic systems. Onlyafter filling these gaps, more reliable environmentalrisk assessments with much lower uncertainty can beperformed.

2. Sources

The consumption of pharmaceuticals is substan-tial. In the European Union (EU) about 3000 differ-ent substances are used in human medicine such asanalgesics and antiinflammatory drugs, contraceptives,antibiotics, beta-blockers, lipid regulators, neuroactivecompounds and many others. Also a large numberof pharmaceuticals are used in veterinary medicine,among them antibiotics and antiinflammatory. Salesfigures are relatively high as reported for several coun-tries (Table 1). In England, Germany and Australia,t arei 2;Hp ffer-e mayb ver,a u-m thec ID)i in2 01),i xen( 6 ti in( car-b rep-r ludem over-

the-counter, some a mixture of both, and internetsales are not included. Therefore, the real amountsof applied drugs is uncertain, but probably signifi-cantly higher for some of the pharmaceuticals reportedthan the figures inTable 1. Figuring out the annualconsumption of a certain drug is difficult and oftenbased on estimates. For example, based on sales, esti-mates of the U.S. production of the antiepileptic car-bamacepine (which is also used for other treatments)ranged from 43 t in 2000 to 35 t in 2003 (Thaker,2005).

Pharmaceuticals are excreted after application intheir native form or as metabolites and enter aquaticsystems via different ways. The main pathway fromhumans is ingestion following excretion and disposalvia wastewater. Municipal wastewater is therefore themain route that brings human pharmaceuticals afternormal use and disposal of unused medicines into theenvironment. Hospital wastewater, wastewater frommanufacturers and landfill leachates (Holm et al., 1995)may contain significant concentrations of pharmaceu-ticals. Pharmaceuticals not readily degraded in thesewage treatment plant (STP) are being dischargedin treated effluents resulting in the contamination ofrivers, lakes, estuaries and rarely, groundwater anddrinking water. Where sewage sludge is applied toagricultural fields, contamination of soil, runoff intosurface water but also drainage may occur. In addi-tion, veterinary pharmaceuticals may enter aquaticsystems via manure application to fields and subse-q ua-c isn cer-t ntalp ight eyf byt ills,s ro-d pery itep enica ar-m e thef ro-d sis-t rme

he amounts for the most frequently used drugsn the hundreds of tons per year (Jones et al., 200uschek et al., 2004; Khan and Ongerth, 2004). Theattern of consumed pharmaceuticals for the dint countries is not identical and some drugse forbidden or replaced by related drugs. Howes listed inTable 1, some drugs are regularly docented within the most frequently applied range:

lass of non-steroidal antiinflammatory drugs (NSAncluding acetylsalicylic acid (e.g. 836 t in Germany001), paracetamol (e.g. 622 t in Germany in 20

buprofen (e.g. 345 t in Germany in 2001), naproe.g. 35 t in England in 2000) and diclofenac (8n Germany in 2001), the oral antidiabetic metforme.g. 517 t in Germany 2001) and the antiepilepticamazepine (e.g. 88 t in Germany 2001). Dataesenting the annual sales or consumptions incainly prescribed drugs, some include also sales

uent runoff, but also via direct application in aqulture (fish farming). Of environmental concernot necessarily a high production volume of a

ain pharmaceutical per se, but the environmeersistence and critical biological activity (e.g. h

oxicity, high potency for effects on biological kunctions such as reproduction). As exemplifiedhe synthetic steroid hormones in contraceptive puch as 17�-ethinylestradiol (EE2), the annual puction lies in a couple of hundreds kilogramsear in the EU, yet it is extremely potent, quersistent in the environment and shows estrogctivity in fish at 1–4 ng/L, or lower. Hence, phaceuticals having environmental relevance shar

ollowing properties: often, but not always, high puction volume combined with environmental per

ence and biological activity, mainly after long-texposure.

126K

.Fentetal./Aquatic

Toxicology76

(2006)122–159

Table 1Annual consumption of different classes of prescribed drugs for different countries

Compounds Germany1999a

Germany2000a

Germany2001a

Austria1997b

Denmark1997c

Australia1998d

England2000e

Italy2001f

Switzerland2004g

Analgesics, antipyretics and anti-inflammatoryAcetylsalicylic acid 902.27 (1) 862.60 (1) 836.26 (1) 78.45 (1) 0.21 (7) 20.4 (9) 43.80 (3)Salicylic acid 89.70 (12) 76.98 (17) 71.67 (17) 9.57 (11) 5.30 (6)Paracetamol 654.42 (2) 641.86 (2) 621.65 (2) 35.08 (2) 0.24 (6) 295.9 (1) 390.9 (1) 95.20 (1)Naproxen 4.63 (16) 22.8 (7) 35.07 (12) 1.70 (12)Ibuprofen 259.85 (5) 300.09 (5) 344.89 (5) 6.7 (13) 0.03 (19) 14.2 (13) 162.2 (3) 1.9 (15) 25.00 (4)Diclofenac 81.79 (16) 82.20 (14) 85.80 (14) 6.14 (15) 26.12 (16) 4.50 (7)

�-BlockerAtenolol 28.98 (13) 22.07 (4) 3.20 (9)Metoprolol 67.66 (18) 79.15 (16) 92.97 (11) 2.44 (20) 3.20 (10)

AntilipidemicGemfibrazol 20 (10) 0.399 (18)Bezafibrate 4.47 (17) 7.60 (8) 0.757 (15)

NeuroactiveCarbamazepine 86.92 (13) 87.71 (13) 87.60 (12) 6.33 (14) 9.97 (18) 40.35 (8) 4.40 (8)Diazepam 0.21 (8) 0.051 (21)

AntiacidicRanitidine 85.41 (15) 89.29 (12) 85.81 (13) 33.7 (5) 36.32 (10) 26.67 (3) 1.60 (13)Cimetidine 35.65 (11) 0.063 (20)

DiureticsFurosemide 3.74 (1) 6.40 (19) 1.00 (14)

SympatomimetikaTerbutalin 0.46 (3) 0.0099 (23)Salbutamol 0.17 (9) 0.035 (22)

VariousMetformin 368.01 (4) 433.46 (4) 516.91 (3) 26.38 (3) 90.9 (2) 205.8 (2) 51.40 (2)Estradiol 0.12 (13)Iopromide 64.93 (19) 63.26 (19) 64.06 (19) 6.90 (5)

For every country a top 20 sold-list is taken into account. Data in bracket represent the position in the ranking list within a country. Data are in t/year.a Huschek et al. (2004).b Sattelberger (1999).c Stuer-Lauridsen et al. (2000).d Khan and Ongerth (2004).e Jones et al. (2002).f Calamari et al. (2003).g ©IMS Health Incorporated or its affiliates. All rights reserved. MIDAS–02/03/05.


3. Fate in the environment

The behavior and fate of pharmaceuticals and theirmetabolites in the aquatic environment is not wellknown. The low volatility of pharmaceuticals indi-cates that distribution in the environment will occurprimarily through aqueous transport, but also via foodchain dispersal. Inwastewater treatment, two elimi-nation processes are generally important: adsorptionto suspended solids (sewage sludge) and biodegrada-tion. Adsorption is dependent on both hydrophobic andelectrostatic interactions of the pharmaceutical withparticulates and microorganisms. Acidic pharmaceu-tical such as the NSAID acetylsalicylic acid, ibupro-fen, fenoprofen, ketoprofen, naproxen, diclofenac andindomethacin having pKa values ranging from 4.9 to4.1, as well as clofibric acid, bezafibrate (pKa 3.6) andgemfibrozil occur as ion at neutral pH, and have littletendency of adsorption to the sludge. But adsorptionincreases with lower pH. At neutral pH, these nega-tively charged pharmaceuticals therefore occur mainlyin the dissolved phase in the wastewater. For these com-pounds and the antitumor agent ifosfamide sorptionby non-specific interactions seems not to be relevant(Kummerer et al., 1997; Buser et al., 1998b). In gen-eral, sorption of acidic pharmaceuticals to sludge issuggested to be not very important for the elimina-tion of pharmaceuticals from wastewater and surfacewater. Therefore, levels of pharmaceuticals in digestedsludge and sediments are suggested to be relativelyl dies(e sorbt ownft ei w-a . Asa herec .,2

thed e them eat-m bic)z lly ins om-p als

increases with increase in hydraulic retention time andwith age of the sludge in the activated sludge treat-ment. For example, diclofenac was shown to be sig-nificantly biodegraded only when the sludge retentiontime was at least 8 days (Kreuzinger et al., 2004). Incontrast, data fromMetcalfe et al. (2003a,b)indicatethat the neutral drug carbamazepine, which is hardlybiodegradable, is only poorly eliminated (normally lessthan 10%), independent from hydraulic retention times.Pharmaceuticals are often excreted mainly as non-conjugated and conjugated polar metabolites. Conju-gates can, however, be cleaved in sewage treatmentplants (STP), resulting in the release of active parentcompound as shown for estradiol (Panter et al., 1999;Ternes et al., 1999), and the steroid hormone in the con-traceptive pill, 17�-ethinylestradiol (D’Ascenzo et al.,2003).

Studies on the elimination rates during the STPprocess are mainly based on measurements of influ-ent and effluent concentrations in STPs, and they varyaccording to the construction and treatment technol-ogy, hydraulic retention time, season and performanceof the STP. Some studies (Ternes, 1998; Stumpf etal., 1999; Carballa et al., 2004) indicate eliminationefficiencies of pharmaceuticals to span a large range(0–99%). The average elimination for specific pharma-ceuticals varied from only 7 to 8% for carbamazepine(Ternes, 1998; Heberer, 2002; Clara et al., 2004) upto 81% for acetylsalicylic acid, 96% for propranolol,and 99% for salicylic acid (Ternes, 1998; Ternes eta alr l ofb eenS xen( tr rma-c ight en,k TPsir eat-m aryt rastm , tota x-i dT is

ow, as was demonstrated in several monitoring stuTernes et al., 2004; Urase and Kikuta, 2005). How-ver, basic pharmaceuticals and zwitterions can ado sludge to a significant extent, as has been shor fluoroquinolone antibiotics (Golet et al., 2002). Forhe hydrophobic EE2 (logKow 4.0) sorption to sludgs likely to play a role in the removal from wasteter. Degradation in sludge seems not significantconsequence, EE2 occurs in digested sludge, w

oncentrations of 17 ng/g were reported (Temes et al002).

In case a pharmaceutical is occurring mainly inissolved phase, biodegradation is suggested to bost important elimination process in wastewater trent. It can occur either in aerobic (and anaero

ones in activated sludge treatment, or anaerobicaewage sludge digestion. In general, biological decosition of micro-pollutants including pharmaceutic

l., 1999; Heberer, 2002). Lowest average removates were found for diclofenac (26%), the removaezafibrate was 51%, but varied significantly betwTPs, and high removal rates were found for napro

81%) (Lindqvist et al., 2005). Table 2 shows thaemoval rates are variable, even for the same phaeutical between different treatment plants. Very hotal elimination of 94–100% of ibuprofen, naproxetoprofen and diclofenac was found in three Sn the U.S.A. (Thomas and Foster, 2004). Efficientemoval took place mainly in the secondary trent step (51–99% removal), whereas in the prim

reatment only 0–44% were removed. X-ray contedia (diatrizoate, iopamidol, iopromide, iomeprol)

he contrary, were not significantly eliminated (Ternesnd Hirsch, 2000). Also, the anticancer drug tamo

fen (antiestrogen) was not eliminated (Roberts anhomas, 2005). This variation in elimination rates


Table 2Influent and effluent concentrations and removal efficiency in sewage treatment plants (different equipment, different countries, sampling indifferent seasons)

Compound Influentconcentration(�g/L)

Effluentconcentration(�g/L)

Maximalremoval(%)

Reference

Analgesics and antiinflammatory drugsAcetylsalicylic acid 3.2 0.6 81 Ternes et al. (1999)

Salicylic acid 57 0.05 99 Metcalfe et al. (2003a)a

330 3.6 Carballa et al. (2004)

Dextropropoxyphene 0.03 0.06 0 Roberts and Thomas (2005)a

Diclofenac 3.0 2.5 17 Heberer (2002)n.r. n.r. 69 Ternes (1998)b

0.33–0.49 n.r. 75 (10–75) Andreozzi et al. (2003a)c

[5] [1.5] 53–74 Strenn et al. (2004)a

1.3 n.r. Metcalfe et al. (2003a)a

0.47–1.9 0.31–0.93 Buser et al. (1998b)2.8 1.9 23± 30 Quintana et al. (2005)b

0.4–1.9 0.4–1.9 0 Tauxe-Wuersch et al. (2005)c

0.35± 0.1 0.17–0.35 9–60 Lindqvist et al. (2005)c

1.0 0.29 71 Roberts and Thomas (2005)a

Ibuprofen 3 96 Buser et al. (1999)38.7 4 >90 Metcalfe et al. (2003a)a

9.5–14.7 0.01–0.02 99 Thomas and Foster (2004)[0.54] [0.08–0.28] 22–75 99 (52–99) Andreozzi et al. (2003a)c

[1.5] [0.01] 12–86 Strenn et al. (2004)a

2.6–5.7 0.9–2.1 60–70 Carballa et al. (2004)a

5.7 0.18 97± 4 Quintana et al. (2005)b


2–3 0.6–0.8 53–79 Tauxe-Wuersch et al. (2005)c

13.1± 4 0–3.8 78–100 Lindqvist et al. (2005)c

Ketoprofen 0.41–0.52 0.008–0.023 98 Thomas and Foster (2004)[0.55] [0.18–0.3] 48–69 Stumpf et al. (1999)b

5.7 n.r. Metcalfe et al. (2003a)a

0.47 0.18 62± 21 Quintana et al. (2005)b

0.25–0.43 0.15–0.24 8–53 Tauxe-Wuersch et al. (2005)c

2.0± 0.6 0–1.25 51–100 Lindqvist et al. (2005)c

Mefenamic acid 1.6–3.2 0.8–2.3 2–50 Tauxe-Wuersch et al. (2005)c


Naproxen 66 Ternes (1998)b

40.7 12.5 40–100 Metcalfe et al. (2003a)10.3–12.8 n.d.-0.023 100 Thomas and Foster (2004)[0.6] [0.1–0.54] 15–78 Stumpf et al. (1999)b

93 (42–93) Andreozzi et al. (2003a)c

1.8–4.6 0.8–2.6 40–55 Carballa et al. (2004)a

0.95 0.27 71± 18 Quintana et al. (2005)b

4.9± 1.7 0.15–1.9 55–98 Lindqvist et al. (2005)c

Paracetamol 6.9 0 100 Roberts and Thomas (2005)a

�-BlockerMetoprolol n.r. n.r. 83 Ternes (1998)b

n.r. n.r. 10 (0–10) Andreozzi et al. (2003a)c


Table 2 (Continued )

Compound Influentconcentration(�g/L)

Effluentconcentration(�g/L)

Maximalremoval(%)

Reference

Propranolol n.r. n.r. 96 Ternes (1998)b

70 304 0 Roberts and Thomas (2005)a

Atenolol n.r. n.r. <10 (0–10) Andreozzi et al. (2003a)c

Blood lipid lowering agentsBezafibrate [1.18] [0.6–0.84] 27–50 Stumpf et al. (1999)b

n.r. n.r. 83 Ternes (1998)b

[5] [0.01] 10–97 Strenn et al. (2004)a

0.6 0.2 Metcalfe et al. (2003a)a

2.6 0.24 91± 4 Quintana et al. (2005)b

0.42± 0.3 0–0.85 15–100 Lindqvist et al. (2005)c

Gemfibrozil n.r. n.r. 69 Ternes (1998)b

[0.3] [0.18–0.28] 16–46 Stumpf et al. (1999)b

n.r. n.r. 75 (10–75) Andreozzi et al. (2003a)c

0.7 1.3 n.r. Metcalfe et al. (2003a)a

Fenofibric acid [0.44] [0.22–0.4] 6–45 Stumpf et al. (1999)b


Clofibric acid n.r. n.r. 6–50 Stumpf et al. (1996)[1] [0.68–0.88] 15–34 Stumpf et al. (1999)b


0.15–0.25 0.15–0.25 0 Tauxe-Wuersch et al. (2005)c

0.34 0 91 Roberts and Thomas (2005)a

Neuroactive compoundsCarbamazepine n.r. n.r. 7–8 Ternes (1998)b

0.7 0.7 <50 Metcalfe et al. (2003a)a

n.r. n.r. 8 Heberer (2002)[1.5] n.r. 4 Clara et al. (2004)a

n.r. [1.5] 53 (0–53) Andreozzi et al. (2003a)c

Diazepam 0.59–1.18 0.1–0.66 93 Van Der Hoeven (2004)

VariousEthinylestradiol 0.003 0.0004 85 Baronti et al. (2000)Clotrimazole 0.031 0.14 55 Roberts and Thomas (2005)a

Ifosfamide 0.007–0.029 0.010–0.043 0 Kummerer et al. (1997)a

Tamoxifen 0.15 0.20 0 Roberts and Thomas (2005)a

X-ray contrast media 0.18–7.5 0.14–8.1 0 Ternes and Hirsch (2000)b

Data estimated from graphical data are in square brackets. n.r.: not reported.a Median concentrations or percent.b Average concentrations or percent.c Maximal concentrations or percent.

not surprising, since pharmaceuticals form a hetero-geneous group consisting of compounds with diversechemical properties. Independent from the chemicalcharacteristics of the compounds, the efficiencies ofvarious STPs also vary for the same compound due totheir equipment and treatment steps but also to other

factors such as temperature and weather. For instance,diclofenac showed largely different elimination ratesbetween 17% (Heberer, 2002) and 69% (Ternes, 1998),and 100% (Thomas and Foster, 2004).

Once insurface waters, biotransformation throughbiodegradation occurs, but abiotic transformation


reactions are probably more important. Whereashydrolysis is generally negligible for environmentallyrelevant human drugs, photodegradation sometimesplays an important role at the water surface. Photolysishas been shown to be the main removal process fordiclofenac in surface water (Buser et al., 1998b).For additional pharmaceuticals (sulfamethoxazole,ofloxacin and propranolol) laboratory experimentsindicate direct and indirect photolysis as an importantremoval process (Andreozzi et al., 2003b). Carba-mazepine and clofibric acid, both compounds thatare marginally processed in STP, have been shown toundergo slow photodegradation in salt- and organic-free water with estimated half-lives in the range of100 days at latitudes of 50◦N in winter (Andreozziet al., 2003b). The efficiency of photodegradationdepends, besides substance properties, on the strengthof the solar irradiation, and therefore on latitudeand season, and on constituents present in the waterthat may act as photosensitizers generating hydroxylradicals and singlet oxygen (i.e. nitrates, humic acids).Some adsorption to particles may occur. Laboratorybatch studies to characterize the sorption behaviorof carbamazepine, diclofenac and ibuprofen in sandysediments show that sorption coefficients were gen-erally quite low (Scheytt et al., 2005). Diclofenac andibuprofen are carboxylic acids with pKa values of 4.16and 4.52 and these weak acids are negatively chargedat pH of ambient water and sediment.

There is no information about thebioaccumulationp itht y ofv dt lser-tb r offi surec

icalsw n,1 se on,a cientr eens ro-f byoi bric

acid, carbamazepine, diclofenac) during drinking watertreatment was investigated in laboratory experimentsand waterworks (Ternes et al., 2002). No significantremoval was observed in batch experiments with sand,indicating low sorption properties and persistence.Flocculation using iron(III) chloride was ineffective,but ozonation was in some cases very effective in elim-inating these polar pharmaceuticals. However, clofibricacid was stable and not eliminated, even with filtrationusing granular activated carbon, which was effectivefor the other compounds. The removal of pharmaceu-ticals and other polar micro-pollutants can thereforeonly be assured using more advanced techniques suchas ozonation, activated carbon or membrane filtration(Ternes et al., 2002). However, the economic conse-quences have to be evaluated carefully before investinginto these advanced treatment technologies on a largerscale.

4. Environmental concentrations

The occurrence of pharmaceuticals was firstreported in the U.S.A. in treated wastewater, whereclofibric acid in the range of 0.8–2�g/L was found(Garrison et al., 1976). Subsequently, pharmaceuticalswere detected in the U.K. in 1981 in rivers up to 1�g/L(Richardson and Bowron, 1985), and ibuprofen andnaproxen were identified in wastewaters in Canada(Rogers et al., 1986). In the last few years, knowledgea ticalsh alyti-c ds att hor-m 7e ale l.,1 to2S 999;KH ntra-t theu ata.F nceo ron-m n thed

otential of pharmaceuticals in biota or food webs whe exception of diclofenac, accumulating in the preultures (Oaks et al., 2004), fluoxetine, sertraline anhe SSRI metabolites norfluoxetine and desmethyraline detected in fish (Brooks et al., 2005). Diclofenacioconcentration factors were 10–2700 in the livesh and 5–1000 in the kidney, depending on expooncentrations (Schwaiger et al., 2004).

A few cases were reported, where pharmaceutere detected indrinking water (Heberer and Sta996) and groundwater (Holm et al., 1995; Ternet al., 2001). Ozonation, granulated activated carbnd advanced oxidation have been shown as effiemoval processes. In drinking water, this has bhown for diclofenac, while clofibric acid and ibupen were oxidized in laboratory experiments mainlyzone/H2O2 (Zwiener and Frimmel, 2000). The elim-

nation of selected compounds (bezafibrate, clofi

bout the environmental occurrence of pharmaceuas increased to a large extent due to new anal techniques able to determine polar compounrace quantities. This also holds for the steroidones contained in contraceptive pills such as 1�-thinylestradiol (EE2), which is linked to biologicffects in fish (Stumpf et al., 1996; Desbrow et a998). Data on environmental concentrations up004 have been compiled and reviewed (e.g.Halling-orensen et al., 1998; Daughton and Ternes, 1ummerer, 2001; Heberer, 2002; Kummerer, 2004).ere, we give a summary on environmental conce

ions focusing on most recent analytical data withltimate aim to relate them to ecotoxicological dirst, we give a general overview on the occurref pharmaceuticals in general and in different enviental media, and subsequently present data oifferent pharmaceutical classes.


Recent studies reported concentrations of a widerange of about 80–100 pharmaceuticals from manyclasses of drugs (antiinflammatory, beta-blockers, sym-pathomimetics, antiepileptics, lipid regulators, antibi-otics, etc.) and some of their metabolites in manycountries in treated sewage, rivers and creeks, sea-water, groundwater and even drinking water.Ternes(1998) reported on the occurrence of 32 pharma-ceuticals belonging to different medicinal classes inGerman municipal STP effluents, river and streamwaters. Twenty different drugs and four correspond-ing metabolites including antiinflammatory drugs(salicylic acid, diclofenac, ibuprofen, indometacine,naproxen, phenazone), lipid regulators (bezafibrate,gemfibrozil, clofibric acid, fenofibric acid), beta-blockers (metoprolol, propranolol) and carbamazepinewere found to be ubiquitously present in streams andriver water in the ng/L range. In an extended monitoringstudy concentrations of 95 micro-pollutants in watersamples of 139 streams downstream of urban areas andlivestock production across the U.S.A. were detected(Kolpin et al., 2002). In some sites as many as 38 ofthe targeted 95 compounds were detected in a singlewater sample (average number of compounds in a sam-ple was seven). Among the most frequently detectedcompounds were steroids (although some data had tobe withdrawn subsequently), an insect repellant (N,N-diethyltoluamide), caffeine, triclosan (an antimicrobialcompound), antibiotics, a fire retardant, 4-nonylphenoland some pharmaceuticals. Analysis of the distribu-t ib-u urg,G ceu-t nac,i lipidr -n ndL gs oticsw amww ectedt al-a iono icalsa amsu ereo s.

Environmental concentrations of pharmaceuticalswere mainly reported in STP effluents and in surfacewater in many countries, often at locations near STPs(Halling-Sorensen et al., 1998; Kolpin et al., 2002;Ashton et al., 2004; Gross et al., 2004). The occur-rence of selected pharmaceuticals was also reportedin the Tyne estuary in the U.K. with concentrationsranging from 4 to 2370 ng/L (Roberts and Thomas,2005). Fig. 1gives a summary on the concentrations ofmost frequently assessed pharmaceuticals in wastewa-ter and surface water reported so far. In STP effluentsa number of different pharmaceuticals occur at con-centrations generally in the ng/L to�g/L range. Inrivers, lakes and seawaters, they are in the ng/L range(Buser et al., 1998b; Kolpin et al., 2002; Weigel et al.,2002; Ashton et al., 2004; Thomas and Hilton, 2004).The rather persistent antiepileptic carbamazepine, andclofibric acid, a metabolite of the lipid lowering agentsclofibrate, etofibrate and etofyllin clofibrate, have beendetected with few exceptions in STP effluents, fresh-water (rivers and lakes) and even in seawater (Buseret al., 1998b; Weigel et al., 2002). In surface water,carbamazepine is found with maximal concentrationsof 1.2�g/L (Wiegel et al., 2004) and clofibric acidat 0.55�g/L (Ternes, 1998). Carbamazepine contam-ination is widespread. In 44 rivers across the U.S.A.average levels were 60 ng/L in water and 4.2 ng/mgin the sediment (Thaker, 2005). Frequently, the anal-gesic ibuprofen and its metabolites were detected inSTP effluents (Ternes, 1998; Buser et al., 1999; Boyde fu r(m vel7 easd %,i ica ene( ian7t undsa

6;S catoe s( ll lsi ,

ion of different drugs in the river Elbe and its trtaries between the source and the city of Hambermany, showed the presence of many pharma

icals. The main substances found were diclofebuprofen, carbamazepine, various antibiotics andegulators (Wiegel et al., 2004). A similar contamiation pattern was found in Italy in the river Po aambro (Calamari et al., 2003) where at all samplinites atenolol, bezafibrate, furosemide, and antibiere found and ranitidine, clofibric acid, diazepere often detected.Kolpin et al. (2004)collectedater samples upstream and downstream of sel

owns and cities in Iowa, U.S.A., during high-, normnd low-flow conditions to determine the contributf urban centres to concentrations of pharmaceutnd other organic wastewater contaminants in strender varying flow conditions. Prescription drugs wnly frequently detected during low-flow condition

t al., 2003; Weigel et al., 2004), in surface water op to 1�g/L (Kolpin et al., 2002), and in seawateThomas and Hilton, 2004; Weigel et al., 2004). In aonitoring study in the U.K. propranolol (median le6 ng/L) was always found in STP effluents, whericlofenac (median 424 ng/L) was found in 86

buprofen (median 3086 ng/L) in 84%, mefenamcid (median 133 ng/L) in 81%, dextropropoxyphmedian 195 ng/L) in 74%, and trimethoprim (med0 ng/L) in 65% of the samples (Ashton et al., 2004). In

he corresponding receiving streams, fewer compond lower levels were found.

Some drinking waters (Heberer and Stan, 199tumpf et al., 1999; Putschew et al., 2000; Zuct al., 2000; Stackelberg et al., 2004), groundwaterHolm et al., 1995; Ternes et al., 2001), and landfileachates (Holm et al., 1995) contain pharmaceutican the ng/L range, in some cases up to�g/L. Phenazone


Fig. 1. Concentration of pharmaceuticals in treated wastewater (a) and surface water (b).References: Halling-Sorensen et al. (1998), Ternes(1998), Stuer-Lauridsen et al. (2000), Jones et al. (2002), Kolpin et al. (2002), Andreozzi et al. (2003b), Calamari et al. (2003), Metcalfe et al.(2003a,b), Gross et al. (2004), Khan and Ongerth (2004), Kummerer (2004), Stackelberg et al. (2004), Thomas and Hilton (2004), Weigel et al.(2004), Lindqvist et al. (2005), Quintana et al. (2005), Roberts and Thomas (2005)andTauxe-Wuersch et al. (2005).

propiphenazone and clofibric acid were found in sam-ples of potable water collected in the vicinity of Berlin,Germany (Heberer and Stan, 1997; Reddersen et al.,2002). Several polar pharmaceuticals such as clofib-ric acid, carbamazepine, and X-ray contrast media canoccur in groundwater. In the following, current knowl-edge about major pharmaceuticals of different thera-peutic classes is summarized.

4.1. Analgesics and antiinflammatory drugs

The widely used non-steroidal antiinflammatorydrugs (NSAID) ibuprofen, naproxen, diclofenac andsome of their metabolites (e.g. hydroxyl-ibuprofen andcarboxy-ibuprofen) are very often detected in sewageand surface water.Ternes (1998)reported levels insewage exceeding 1�g/L, and in effluents of con-ventional STP (mechanical clarification and biologicaltreatment) concentrations often approach or exceed

0.1�g/L in the U.S.A. (Gross et al., 2004). The deacy-lated, more active form of acetylsalicylic acid, salicylicacid, has been found in many municipal wastewatersat levels up to 4.1�g/L (Ternes, 1998), 13�g/L (Farreet al., 2001; Heberer, 2002) or even 59.6�g/L withmedian levels of 3.6�g/L (Metcalfe et al., 2003a).However, salicylic acid may also derive from othersources. Similar to acetylsalicylic acid, acetaminophen(paracetamol) is well removed from STP. However,up to 10�g/L (median 0.11�g/L) acetaminophen hasbeen found in 24% of samples from U.S. streams(Kolpin et al., 2002). The analgesic codeine wasdetected in 7% of samples at median concentrationsof 0.01�g/L.

In many countries diclofenac was frequentlydetected in wastewater in the�g/L range, and in sur-face water at lower levels (Heberer and Stan, 1997;Buser et al., 1998b; Ternes, 1998; Stumpf et al., 1999;Farre et al., 2001; Sedlak and Pinkston, 2001; Heberer,


2002). This also holds for ibuprofen (Heberer and Stan,1997; Ternes, 1998; Buser et al., 1999; Stumpf etal., 1999). Sometimes, high levels of up to 85�g/L(Farre et al., 2001), or 24.6�g/L (median 4.0�g/L)were detected in STP effluents (Metcalfe et al., 2003a).In Norway, ibuprofen and its metabolites occurred inall sewage samples, and in seawater at concentrationsof 0.1–20�g/L (sum of ibuprofen and metabolites)(Weigel et al., 2004). In U.K. estuaries maximal con-centration of 0.93�g/L (median 0.05�g/L) occurred(Thomas and Hilton, 2004). Ibuprofen is significantlyremoved during sewage treatment, and metabolitessuch as hydroxy-ibuprofen occur in STP effluents.Kolpin et al. (2002)found ibuprofen in 10% of streamwater samples with maximal concentrations of 1�g/L(median 0.2�g/L). In two stormwater canals levels ofibuprofen were up to 674 ng/L and of naproxen up to145 ng/L (Boyd et al., 2004). Naproxen was also foundat much higher level in Canadian STP effluents withmedian levels of 12.5�g/L and maximal levels of upto 33.9�g/L (Metcalfe et al., 2003a). Moreover, sev-eral other analgesics have been detected in sewage andsurface water, but also in ground water and drinkingwater samples.

4.2. Beta-blockers

Several beta-blockers were identified in wastewater(Ternes, 1998; Sedlak and Pinkston, 2001). Propra-nolol, bisoprolol and metoprolol were found at highestl ew er)a ,1 vea und-w

4

s ofw llinc ntlyf tud-i rfacew l.,1 ra-t 7a ,

1996; Heberer and Stan, 1997). Bezafibrate occurredin maximal concentrations of up to 4.6 and 3.1�g/L(median 2.2 and 0.35�g/L, respectively) in wastewa-ter and surface water, respectively (Stumpf et al., 1996;Ternes, 1998). In addition, gemfibrozil, clofibric acidand fenofibric acid (metabolite of fenofibrate) have alsobeen detected in sewage up to the�g/L level and in sur-face water (Ternes, 1998; Stumpf et al., 1999; Farre etal., 2001; Heberer, 2002). Gemfibrozil was detected in4% of streams at maximal levels of 0.79�g/L (Kolpinet al., 2002).

4.4. Neuroactive compounds (antiepileptics,antidepressants)

Of this category, the antiepileptic carbamazepinewas detected most frequently and in highest concen-tration in wastewater (up to 6.3�g/L) (Ternes, 1998),and at lower levels in other media (Heberer et al.,2002; Andreozzi et al., 2003b; Metcalfe et al., 2003b;Wiegel et al., 2004). Carbamazepine was found in everyCanadian STP effluent sample at concentration up to2.3�g/L (Metcalfe et al., 2003b). This compound wasfound to occur ubiquitously in the river Elbe and itstributaries, Germany (Wiegel et al., 2004), exceed-ing 1�g/L in other German surface waters (Ternes,1998; Heberer, 2002) and occurred in groundwater(Seiler et al., 1999; Sacher et al., 2001; Ternes et al.,2001). In U.S. rivers average levels were 60 ng/L inwater and 4.2 ng/mg in the sediment (Thaker, 2005).C ls of2 t in8 en-tB( new nada( ance asa ,2

4

occurp n-c and

evels (0.59, 2.9 and 2.2�g/L, respectively, in surfacater), with lower levels of nadolol (in surface watnd betaxolol (0.028�g/L in surface water) (Ternes998). Propranolol, metoprolol and bisoprolol halso been found in surface water, and sotalol in groater (Sacher et al., 2001).

.3. Blood lipid lowering agents

Clofibric acid, the active metabolite from a serieidely used blood lipid regulators (clofibrate, etofylofibrate, etofibrate) belongs to the most frequeound and reported pharmaceutical in monitoring ses. It has been found in numerous wastewaters, suaters, in seawater (Stumpf et al., 1996; Buser et a998a; Ternes, 1998), and at rather high concent

ions in groundwater (4�g/L) (Heberer and Stan, 199)nd drinking water (0.07–0.27�g/L) (Stumpf et al.

arbamazepine was also found at average leve0.9 ng/mg solids of STP. Diazepam was presenof 20 STPs in Germany at relatively low conc

rations of up to 0.04�g/L (Ternes, 1998) whereas inelgium it was found at concentration up to 0.66�g/L

van der Ven et al., 2004). The antidepressant fluoxetias also detected in STP effluents samples in Ca

Metcalfe et al., 2003a), and in U.S. streams, medioncentrations of 0.012�g/L were estimated (Kolpint al., 2002). Primidone, an antiepileptic drug, hlso been detected in sewage up to 0.6�g/L (Heberer002).

.5. Antineoplastics and antitumor agents

Pharmaceuticals used in cancer chemotherapyrimarily in hospital effluents and only at lower coentrations in municipal wastewater. Ifosfamide


cyclophosphamide occur in concentrations of up to4.5�g/L in hospital wastewaters (Steger-Hartmannet al., 1997), and at ng/L in municipal wastewater(Kummerer et al., 1997; Steger-Hartmann et al., 1997).The occurrence of the antiestrogen tamoxifen used inbreast cancer therapy was reported in U.K. wastewater,where concentrations in STP effluents ranged between146 and 369 ng/L (Roberts and Thomas, 2005). Tamox-ifen was not reduced in the STP, and even foundin estuarine waters (Tye estuary) at concentrationsranging from 27 to 212 ng/L with a median levelof 53 ng/L (Thomas and Hilton, 2004; Roberts andThomas, 2005).

4.6. Various other compounds

Many additional pharmaceuticals have beendetected in sewage and surface water (Daughton andTernes, 1999; Heberer, 2002; Kolpin et al., 2002).Here only a few of them will be mentioned. Thestimulant caffeine and the nicotine metabolite cotininewere generally present in STP effluents and surfacewaters contaminated by drugs (Metcalfe et al., 2003b).Caffeine was generally found in U.S. streams atmaximal levels of 6.0�g/L (median 0.1�g/L) (Kolpinet al., 2002) and this compound can even serve asan anthropogenic marker in aquatic systems due toits ubiquity in surface water, seawater (Weigel et al.,2004), and also in groundwater (Fig. 1). The antiacidcimetidine and ranitidine were estimated to occur inUr cy of1c beenfsew to2d eta rceoa in5 velso s( ol)w note

4.7. Steroidal hormones

Steroidal hormones have been reported on in manyreports, and in our review we only summarize knowl-edge about the synthetic estrogen EE2 and mestranolcontained in contraceptive pills. These steroids havebeen found in numerous studies in many countries inEurope, Canada, the U.S.A., Japan, Brazil, etc. both inwastewater and surface water. A survey in the U.S.A.showed that maximal and median EE2 concentrationswere as high as 831 and 73 ng/L, respectively, and lev-els of mestranol were 407 and 74 ng/L, respectively(Kolpin et al., 2002). They were detectable in 16 and10% of the streams sampled. Generally, median con-centrations are much lower being in the range of non-detectable up to 9 ng/L in treated wastewater in severalcountries (Baronti et al., 2000). Typical wastewatereffluent concentrations are 0.5 ng/L and they are evenlower in surface water. However, these concentrationsmust put into the perspective of their high biologicalactivity accounting for potential estrogenic effects infish.

Exposure and fate models are increasingly beingused to estimate environmental concentrations with-out analytical chemical measurements. Some exposuremodels have been developed for drugs (e.g. PhATE),others have been extended from general chemicals topharmaceuticals (e.g. EPIWIN, GREAT-ER). Thesetools have been developed both for estimation of pre-dicted environmental concentrations (PEC) and theb t. Ap ationm cen-t .S.sm theU sur-f d forp cedR eanR co-l riverbG ofP , fori ver-a usesa al

.S. streams at concentrations of 0.58 and 0.01�g/L,espectively, and they were detected at a frequen0 and 1%, respectively (Kolpin et al., 2002). X-rayontrast media are very persistent. Iopamidol hasound in municipal wastewater as high as 15�g/L, inurface water (0.49�g/L) and groundwater (Putschewt al., 2000; Ternes and Hirsch, 2000). Iopromideas detected at 2–4�g/L in surface water, and up1�g/L in STP (Putschew et al., 2000), but showedegradation in the laboratory (Steger-Hartmannl., 2002). Hospital wastewater was also a souf gadolinium (Kummerer and Helmers, 2000). Thentidiabetic compound metformin was observed% of stream water samples with estimated lef 0.11�g/L (Kolpin et al., 2002). Bronchodilator�2-sympathomimetics terbutalin and salbutamere also detected in sewage in a few casesxceeding 0.2�g/L (Ternes, 1998).

ehavior of pharmaceuticals in the environmenharmaceutical assessment and transport evaluodel (PhATE) was developed to estimate con

rations of active pharmaceutical ingredients in Uurface waters (Anderson et al., 2004). The PhATEodel uses some for most hydrologic regions of.S. representative watersheds. For European

ace waters an exposure simulation was developeharmaceuticals with the GREAT-ER (Geo-referenegional Exposure Assessment Tool for Europivers) model, a tool developed for use within e

ogical risk assessment (ERA) schemes andasin management (Schowanek and Webb, 2002). TheREAT-ER software calculates the distributionEC’s of consumer chemicals in surface waters

ndividual stretches, as well as representative age PEC’s for entire catchments. The systemn ARC/INFO-ArcView (ESRI®) based Geographic


Information System (GIS) for data storage and visu-alization, combined with simple mathematical modelsfor prediction of the fate of chemicals.

For some estimates, measured environmental con-centrations (MEC) are in agreement with the estimatedPEC’s, however, often, they are not as large differ-ences occur between the models and the real worldsituation. The main reason is that different assump-tions are made, which not always correspond to the realconditions in the environment. Consumption figures,metabolism in the organism, removal during sewagetreatment plants and fate in the environment containall uncertainties that may result in inappropriate esti-mates of PEC’s. Moreover, detailed situations at a givensite is not reflected by models integrating large geo-graphical areas. Poor prediction performance of currentmodels for many pharmaceuticals is one of the out-standing scientific issues with regard to the questionof pharmaceuticals in the environment. It is hoped thatthe models are improving by further refining the men-tioned uncertainties and may be developing to a usefuland readily applicable regulatory tool (Sanderson et al.,2004b).

5. Modes of actions in humans and mammalsand occurrence of target biomolecules in lowervertebrates and invertebrates

Here, we briefly summarize the modes of actionso sim-i werv simi-l ttlei rgetb . Ina dis-c wera

5a

byi oft OX-1 dif-f

Botting, 1998). Classical NSAID inhibit both COX-1and COX-2 at different degrees, whereas new NSAIDact more selectively on COX-2, the inducible formresponsible for the inflammatory reactions. Differencesin binding site size are responsible for the selectivityof these drugs (Kurumbail et al., 1997; Penning et al.,1997; Gierse et al., 1999). NSAIDS are commonly usedto treat inflammation and pain and to relieve fever, andsometimes they are also used for long-term treatmentof rheumatic diseases.

Prostaglandins play a variety of physiologicalroles according to their cells source and targetmolecules. They are known to be involved in pro-cess such as inflammation and pain, regulation ofblood flow in kidney, coagulation processes and syn-thesis of protective gastric mucosa (Smith, 1971; Vane,1971; Mutschler, 1996). Since NSAID inhibit non-specifically prostaglandin synthesis, most side effects,at least after long-term treatment, are related to thephysiological function of prostaglandins. In the kid-ney, prostaglandins are involved in maintenance of theequilibrium between vasoconstriction and vasodilata-tion of the blood vessel that supply glomerular filtra-tion. Renal damages and renal failure after chronicNSAID treatment seems to be triggered by the lackof prostaglandins in vasodilatation-induction. Gastricdamages are thought to be caused by inhibition of bothCOX isoforms (Wallace, 1997; Wallace et al., 2000). Incontrast, liver damages are apparently due to buildingof reactive metabolites (e.g. acyl glucuronides) rathert1

yetf inlyb er-v torye ph-e ro-c ainlyd ar-i ent erc al-g ofc with-o y inr -s

f pharmaceutical classes and ask, whether or notlar target receptors and biomolecules exist in loertebrates and invertebrates. Knowledge aboutar targets exists primarily for fish. In general, very lis known about possible counterparts of human taiomolecules of pharmaceuticals in invertebratesddition, some of the side effects in humans areussed, giving hints to possible adverse effects in lonimals.

.1. Analgesics and non-steroidalntiinflammatory drugs (NSAID)

Non-steroidal antiinflammatory drugs actnhibiting either reversibly or irreversibly one or bothhe two isoforms of the cyclooxygenase enzyme (C

and COX-2), which catalyze the synthesis oferent prostaglandins from arachidonic acid (Vane and

han inhibition of prostaglandins synthesis (Bjorkman,998).

The mode of action of paracetamol is notully elucidated. It seems that this drugs acts may inhibiting the cyclooxygenase of the central nous system and it does not have antiinflammaffects, because of the lack of inhibition of periral cyclooxygenase involved in inflammatory pesses. Adverse effects of paracetamol are mue to formation of hepatotoxic metabolites, prim

ly N-acetyl-p-benzoquinone imine, synthesized whhe availability of glutathione is diminished in livells. Acetaminophen widely used in many anesic/antipyretic medications induces proliferationultured breast cancer cells via estrogen receptorsut binding to them, but has no estrogenic activitodents (Harnagea-Theophilus et al., 1999). The conequences of these observations are not clear.


In fish an inducible COX-2 homologue has beenfound to be expressed in macrophages in rainbow trout(Oncorhynchus mykiss) and the translation product ofthe COX gene was found to have a high homology of83–84 and 77% to its human counterpart COX-2 andCOX-1, respectively (Zou et al., 1999). Also in gold-fish, macrophages express a COX enzyme, which is anequivalent to mammalian COX-2 (Zou et al., 1999).A COX-1 and COX-2 homologue was cloned frombrook trout ovary (Roberts et al., 2000), and recently,a shark COX was cloned in dogfishSqualus acan-thias having 68 and 64% homology to mammalianCOX-1 and COX-2, respectively (Yang and Carlson,2004). Prostaglandins are formed in a diverse rangeof vertebrates and invertebrates. However, in lowerinvertebrates such as corals, their synthesis is inde-pendent of COX, involving other enzymes (Song andBrash, 1991). In arthropods and molluscs, COX-likeactivity is apparently responsible for the formationof prostaglandins, but these enzymes have not beenpurified and characterized (Pedibhotla et al., 1995).In birds, prostaglandins play a role in the biosynthe-sis of egg shells and treatment with the COX-inhibitorindometacine resulted in egg shell thinning (Lundholm,1997).

5.2. Beta-blockers

Beta-blocker act by competitive inhibiting beta-adrenergic receptors and they are used in the treatmento reatp . Thea icalf eeda ves-s ergics andl eedss

ptorp lti-m engesi velyib rdiace theh ical

groups added to compounds that are able to enhancethe interactions with amino acids of the transmem-brane domains. Some of the beta-blockers (e.g. pro-pranolol, a beta1-adrencoceptor antagonist) have theability to cause cell membrane stabilization, whereasother (e.g. metoprolol) have no membrane stabilizingactivity (Doggrell, 1990). Side effects of this therapeu-tic class are mainly bronchoconstriction and disturbedperipheral circulations (Hoffman and Lefkowitz, 1998;Scholze, 1999). Due to their lipophilicity they are sup-posed to pass the blood brain barrier and to act in thecentral nervous system (Soyka, 1984, 1985).

�-Adrenoceptors were found in fish (O. mykiss)liver, red and white muscle with a high degree ofsequence conservation with other vertebrate homo-logues. They are also supposed to play similar roleas in humans (Nickerson et al., 2001). The presenceof a �2-adrenoceptor subtype was also suggested bybinding studies to occur in liver membranes of otherfish and amphibians.�2-Adrenoceptors of rainbowtrout (Nickerson et al., 2001) show a high degree ofamino-acid sequence conservation with other verte-brate�2-adrenoceptors. Frog- (Devic et al., 1997) andturkey�1-adrenoceptors (Yardeny et al., 1986) are sim-ilar to mammalian�1-adrenoceptors. In rainbow trout,the �2-adrenoceptor gene is highly expressed in theliver, red and white muscle, with lower expression ingills, heart, kidney and spleen (Nickerson et al., 2001).Clenbuterol or ractopamine that function in mammalsas �-agonist were found in rainbow trout to showa yedo ofr out� ep-a ofte iono s int redb

s,fi undei wsm als.T fisha andl ific

f high blood pressure (hypertension), and to tatients after heart attack to prevent further attacksdrenergic system is involved in many physiolog

unctions such as regulation of the heart oxygen nnd beating, vasodilatation mechanisms of bloodels, and bronchodilation. Furthermore, the adrenystem is also known to interact with carbohydrateipid metabolisms, mainly in response to stress nuch as starvation (Jacob et al., 1998).

�-Adrenoceptors are 7-transmembrane receroteins coupled with different G-proteins that uately enhance the synthesis of the second mess

ignaling molecules cAMP (Rang et al., 2003). Accord-ng to medical needs beta-blockers may selectinhibit one or more�-receptors types; for example�2-lockers are used to treat hypertension avoiding caffects, since this receptor subtype is not found ineart. Selectivity is based on difference in chem

r

somewhat different reaction. Clenbuterol displanly partial agonist activities and the small effectsactopamine may be related to low affinity for the tr2-adrenoceptor. Agonist regulation of the trout htic�2-adrenoceptors may involve down-regulation

he receptors without affecting responsiveness (Dugant al., 2003). Differences in the structure and functf the receptors may be responsible for difference

he affinity with�-blockers and mechanisms triggey these drugs.

Whereas mammals have three�2-adrenoceptorve distinct �2-adrenoceptor genes have been foxpressed in zebrafish (Ruuskanen et al., 2005). Local-

zation of the �-adrenoceptors in zebrafish shoarked conservation when compared with mammhe �2-adrenergic system is functional in zebras demonstrated by marked locomotor inhibition

ightening of skin color induced by the spec


�2-adrenoceptor agonist dexmedetomidine, similar tomammals. Both effects were antagonized by the spe-cific �2-adrenoceptor antagonist atipamezole. The�-adrenoceptor agonists medetomidine and clonidineare being investigated as potential antifouling agentspreventing the settlement of barnacles on ship halls(Dahlstrom et al., 2004). Settlement of larvae is inhib-ited at low concentrations of 0.25–2.5�g/L. Additionalpharmacological and biochemical investigations on�-and�-adrenoceptors of fish and other lower organismsare needed.

5.3. Blood lipid lowering agents

There are basically two types of antilipidemic drugs,namely statins and fibrates, the latter have been tar-geted analytically more often in the aquatic environ-ment than the former. Both types are used to decreasethe concentration of cholesterol (statins and fibrates)and triglycerides (fibrates) in the blood plasma. Statinsas inhibitors of cholesterol synthesis act by inhibit-ing the 3-hydroxymethylglutaril coenzyme A reduc-tase (HMG-CoA), responsible for the limiting step inthe cholesterol synthesis, namely the conversion ofHMG-CoA to mevalonate (Laufs and Liao, 1998). Asa consequence of the intracellular cholesterol deple-tion, the expression of LDL receptors in hepatocytemembranes is increased and therefore, the resorptionof LDL-cholesterol from blood plasma. Due to interac-tions of statins with mevalonate metabolism, multiplea da-t nileha sis inva

, atl ofg einm thel on-s ro-t ),d ntra-t i-s learr llu-l lipid

regulatory proteins such as, for example, the lipopro-tein lipase (Staels et al., 1998). To date, three subtypesof PPAR have been described; PPAR� is involved inperoxisome proliferation and plays a pivotal role incontrolling hepatic lipid metabolism (Schoonjans etal., 1996), whereas PPAR� has diverse roles in basiclipid metabolism, and PPAR� plays a key role in thedifferentiation of adipocytes (Kersten et al., 2000). Het-erodimerization of PPARs with the retinoid X receptorand their binding to response elements in the promoterregions of genes leads to their activation.

Fibrates stimulate cellular fatty acid uptake, con-version to acetyl-CoA derivatives, and catabolism bythe beta-oxidation pathways, which, combined witha reduction in fatty acid and triglyceride synthesis,results in a decrease in VLDL production (Staels et al.,1998). Hepatic damages may occur after chronic expo-sure to fibrates in rat (Qu et al., 2001) and this is thoughtto be related to inhibition of mitochondrial oxidativephosphorylation (Keller et al., 1992). Furthermore,fibrates caused in rodents a massive proliferation ofperoxisomes (Hess et al., 1965). Strong correlationbetween fibrates exposure and hepatocarcinogenicityin rodents were found, while this was not observedin humans (Cajaraville et al., 2003). These findingsincrease the interest for ecotoxicological impact of thistherapeutic class of drugs.

PPAR genes have been found in fish such as plaice(Leaver et al., 1998) and Atlantic salmon (Ruyter et al.,1997) and zebrafish (Ibabe et al., 2002). Fish PPARsd 8%t .,2 sh,a andt cids(t brateit Ror in thel erntA e-n rox-y rates( ef

dditional effects occur (antiinflammatory, antioxiive). There is also evidence that statins affect juveormone synthesis in insects (Debernard et al., 1994),s fluvastatin completely suppressed its biosyntheitro, and in the mandibular organo of lobsters (Li etl., 2003).

In contrast, effects of fibrates are mediatedeast in part, through alterations in transcriptionenes encoding for proteins controlling lipoprotetabolism. Fibrates act probably by activating

ipoprotein lipase enzyme, which is mainly respible for the conversion of very low density lipopein (VLDL) to high density lipoproteins (HDLecreasing therefore plasma triglycerides conce

ion (Staels et al., 1998). Binding of fibrates to peroxome proliferator-activated receptors (PPARs), nuceceptors known to be activated during different cear pathways, stimulates the expression of several

isplay an amino acid sequence identity of 43–4o the human and amphibian PPAR� (Andersen et al000). All PPAR forms have been found in zebrafind PPAR� was mainly expressed in hepatocyte

issues that catabolize high amounts of fatty aIbabe et al., 2002). Furthermore, PPAR� was showno be induced in response to clofibrate and bezafin salmon hepatocytes (Ruyter et al., 1997), althoughheir PPAR� seem to be less responsive than PPA�f rodents (Andersen et al., 2000). All three PPAReceptors were found to already been expressedarval stage, with a similar tissue distribution patto that found in adult zebrafish (Ibabe et al., 2005a).ctivators of PPAR� include a variety of endogously present fatty acids, leukotrienes and hydeicosatetraenoic acids and drugs, such as fibCajaraville et al., 2003). PPAR� activators includatty acids, prostaglandin A2 and prostacyclin. PPAR�


is the most selective receptor and prostaglandin J2 hasbeen described to be a specific ligand (Ibabe et al.,2005b). In isolated zebrafish hepatocytes, mRNA ofboth PPAR� and PPAR� was induced by clofibrateat 0.5–2 mM, although to a low extent (Ibabe et al.,2005b). The physiological and toxicological roles ofPPARs have yet to be investigated, and their involve-ment in potential effects of lipid lowering drugs is notyet known. With regard to invertebrates, no informa-tion is currently available on the existence of PPARs,although extensive searches for nuclear receptors incnidarians and platyhelminthes have been performed(Escriva et al., 1997).

5.4. Neuroactive compounds (antiepileptics,antidepressants)

Among the many drugs interacting with the cen-tral nerve system (CNS), only a few will be consid-ered as with respect to its occurrence in the aquaticenvironment. Antiepileptic drugs act on the CNS bydecreasing the overall neuronal activity. This can beachieved either by blocking voltage-dependent sodiumchannels of excitatory neurons (e.g. carbamazepine),or by enhancing of inhibitory effects of the GABAneurotransmitter by binding on a specific site in thegamma subunit of the corresponding receptor (e.g.diazepam, member of benzodiazepine family) (Studyand Barker, 1981; MacDonald and Olsen, 1994; Rogerset al., 1994). Evidence of the occurrence of the GABAs dE aveb olt-a te.

hicha eu-r elyh t inf r. Ap backt takei thes as an ver-t tedw iblyt oth-e h as

fingernail claims (Sphaerium striatinum, Fong et al.,1998) and Japanese medaka (Oryzias latipes, Fonget al., 1998; Foran et al., 2004). Fluoxetine and ser-traline and the SSRI metabolites norfluoxetine anddesmethylsertraline have been detected in fish sampledfrom wild in the U.S., and therefore reflect a bioaccu-mulation potential (Brooks et al., 2005). Whether theaccumulated levels of 1.6 ng/g fluoxetine and 4.3 ng/gsertraline found in brain have effects on the nervoussystem of fish has yet to be investigated.

5.5. Cytostatics compounds and cancertherapeutics

Another potential interesting class of compoundis represented by cytostatic pharmaceuticals interact-ing with cell proliferation. There are different modesof actions of the different compounds. For examplemethotrexate acts as a potent inhibitor of the folatedehydroreductase enzyme, which is responsible forthe purine and pyrimidine synthesis (Schalhorn, 1995;Rang et al., 2003). Doxorubicin is an intercalating sub-stance inducing DNA-strand brakes (in humans, heartarrhythmia may be a side effect). Tamoxifen as an antie-strogenic drug is used for breast cancer treatment andacts by competitive inhibiting the estrogenic receptorat least in mammary gland (Rang et al., 2003).

5.6. Various compounds

actb as-t id).T inceH bothd ands )d nct asa mayo Hr

ha-n thatt cosea emst the

ystem in fish (O. mykiss, Cole et al., 1984; Meissl ankstrom, 1991) was found, whereas no studies heen found indicating the occurrence of sodium vge dependent channels in fish or lower invertebra

Fluoxetine is a widely used antidepressant, wcts by inhibiting the re-uptake of serotonin. This notransmitter is involved in many mechanisms, namormonal and neuronal, and it is also importan

unctions such as food intake and sexual behavioump directs serotonin from the synapse space

o the presynapse, and selective serotonin re-upnhibitors (SSRI) inhibit this pump, thus increasingerotonin level in the synapse space. Serotonineurotransmitter occurs in lower vertebrates and in

ebrates (Fong, 1998), however, the effects associaith this transmitter are different, and so are poss

he effects of SSRI. Serotonin mediates, amongrs, endocrine functions in aquatic organisms suc

Cimetidine and ranitidine are compounds, whichy inhibiting the histamine receptors type 2 in the g

ric system, thus inhibiting the acid secretion (antachese drugs are used to treat gastric ulceration. S2-histamine receptors are found also in the brain,rugs may elicit central nervous system reactionside effects (Cannon et al., 2004). Peitsaro et al. (2000emonstrated the presence of H3-histamin receptors ientral nervous system of zebrafish (Danio rerio), buthe lack of histamine in the periphery of this fish wlso reported. However, interspecies differencesccur; cod and carp seem to have histamine and2-eceptors in the periphery (Peitsaro et al., 2000).

Metformin is an antidiabetic agent, which mecisms of actions are not well understood. It seems

his drugs acts by increasing the cellular use of glund inhibiting the gluconeogenesis. Metformin se

o act on insulin receptor by direct stimulation of


insulin receptor or indirectly through inhibition of tyro-sine phosphatase (Holland et al., 2004).

6. Ecotoxicological effects

Pharmaceuticals are designed to target specificmetabolic and molecular pathways in humans and ani-mals, but they often have important side effects too.When introduced into the environment they may affectthe same pathways in animals having identical or sim-ilar target organs, tissues, cells or biomolecules. Asshown above, certain receptors in lower animals resem-ble those in humans, others however, are different orlacking, which means that dissimilar modes of actionsmay occur in lower animals. It is important in thisrespect to recognize that for many drugs, their specificmodes of actions are not well known and often not onlyone, but many different modes of actions occur. Amongother reasons, this makes specific toxicity analysis inlower animals difficult to perform. Despite this, toxi-city experiments should be targeted and designed forspecific targets of the pharmaceutical even in lower ver-tebrates and invertebrates, based on the hypothesis ofsimilarity of modes of actions. However, current toxi-city testing is not designed in this way, rather generaland established test systems and traditional organismsaccording to guidelines are being used and traditionalend points such as mortality are assessed.

Thus far, ecotoxicity testing merely providedi ofd re,a res.T riska reen-i orei thet s ab yi evel-o cts.T ivityr ineh( kss rma-c ed toe t

al., 2002; Sanderson et al., 2004b; Cleuvers, 2005).Both methods are helpful in estimating potential toxic-ity or the behavior of a compound in the environment,but they cannot replace in vivo or in vitro assays.

The current literature about ecotoxicological effectsof human pharmaceutical deals mainly with the acutetoxicity in standardized tests and it is generally focusedon aquatic organisms. The influence of environmen-tal parameters such as pH on toxicity has only rarely,or not yet been investigated. Such studies would beof importance for instance for acidic pharmaceuticalsthat may induce different toxicities depending on spe-ciation at different ambient pH. Moreover, effects ofdrug metabolites have rarely been investigated. Pho-totransformation products of naproxen, for instance,showed higher toxicities than the parent compound,while genotoxicity was not found (Isidori et al., 2005).At contaminated sites, aquatic life is exposed over theentire life cycle to these compounds. Chronic effectsare less investigated and often even related to relativeshort-term exposures. However, long-term exposuresare needed for an accurate environmental risk assess-ment. Here we summarize the current ecotoxicologicaldata, focusing on specific modes of action of differ-ent therapeutic classes of pharmaceuticals, and cover-ing many differences in methods, species and time ofexposure. These data are then related to environmentallevels in order to assess the potential hazard for the dif-ferent classes of pharmaceuticals and identify currentresearch and knowledge gaps.

6

xicityb shedg tab-l ktona a ofp ne al rents els,W itivetfi eral-i deso dif-f pt

ndications of acute effects in vivo in organismsifferent trophic levels after short-term exposund only rarely after long-term (chronic) exposuhese data are ultimately used for ecologicalssessments. Because of animal welfare and sc

ng purposes, in vitro analyses are becoming mmportant, but they are not sufficient for assessingoxicological profiles of a compound, particularly aasis for risk analysis (Fent, 2001). Beyond laborator

nvestigations, some mathematical models were dped to estimate or predict ecotoxicological effehe most often applied quantitative structure–actelationship (QSAR) program is ECOSAR (onlttp://www.epa.gov/oppt/newchems/sarman.pdf)Sanderson et al., 2004b). Despite serious drawbacuch as an inadequate structure coverage for phaeuticals, the program has been repeatedly applistimate pharmaceutical baseline toxicities (Jones e

.1. Acute effects

Pharmaceuticals are assessed for their acute toy traditional standard tests according to establiuidelines (e.g. OECD, U.S. EPA, ISO) using es

ished laboratory organisms such as algae, zooplannd other invertebrates and fish. Acute toxicity datharmaceuticals were compiled byHalling-Sorenset al. (1998)andWebb (2001), whereby in the latter,

ist of about 100 human pharmaceuticals from diffeources is given. By comparing different trophic levebb (2001)suggested that algae were more sens

o the listed pharmaceuticals thanDaphnia magna, andsh were even less sensitive. However, such genzations do not focus enough on the different mof actions of a given pharmaceutical, and hence,

erences in toxicity in different phyla. In the attem

http://www.epa.gov/oppt/newchems/sarman.pdf


to compare the different classes of pharmaceuticals interms of acute toxicity,Webb (2001)noted that themost toxic classes were antidepressants, antibacterialsand antipsychotics, but the range of responses withineach of these categories was large, typically severalorders of magnitude. In our present review, we provideand summarize additional and new data and discuss itsecotoxicological relevance covering different classesof human pharmaceuticals. The data originate fromdifferent sources, and studies were performed underdifferent quality criteria (i.e. nominal versus mea-sured exposure concentrations), making comparisonsdifficult.

6.1.1. Analgesics and non-steroidalantiinflammatory drugs (NSAID)

In general, toxicity data vary for each pharma-ceutical, however, diclofenac seems to be the com-pound having highest acute toxicity within the classof NSAID, since for all the tests performed the effect

concentrations were below 100 mg/L (Fig. 2). Short-term acute toxicity was analyzed in algae and inver-tebrates (Webb, 2001; Cleuvers, 2003), phytoplank-ton was found to react more sensitive [lowest EC50(96 h) = 14.5 mg/L (Ferrari et al., 2004)] than zooplank-ton [lowest EC50 (96 h) = 22.43 mg/L (Ferrari et al.,2004)]. There is no correlation between the acute toxi-city in Daphnia and the lipophilicity as represented bylogKow (Fig. 3). In general, not much is known aboutthe acute toxicity to fish.

6.1.2. Beta-blockersAs shown in Fig. 2, the acute toxicity of beta-

blockers is not extensively studied, with the excep-tion of propranolol. This compound shows the high-est acute toxicity and highest logKow as compared toother beta-blockers (Fig. 3). This and the fact that itis a strong membrane stabilizer, whereas other inves-tigated beta-blockers are not, may in part explain itshigher toxicity (Doggrell, 1990; Huggett et al., 2002).

F to diffeo arized( l. (199,( nry et l.(

ig. 2. Acute toxicity of 24 different pharmaceuticals, belongingrganisms and different endpoint and exposure time are summ1994), Henschel et al. (1997), Fong (1998), Halling-Sorensen et a2003, 2004), Villegas-Navarro et al. (2003), Ferrari et al. (2004), He2004a,b), Nunes et al. (2004)andIsidori et al. (2005).

rent therapeutic classes to aquatic organisms. EC50 and LC50 for different. See text for details.References: Calleja et al. (1993, 1994), Lilius et al.

8)Webb (2001), Huggett et al. (2002), Brooks et al. (2003), Cleuversal. (2004), Hernando et al. (2004), Kummerer (2004), Marques et a


Fig. 3. Relation between acute toxicity (LC50) of analgesics and anti-inflammatory drugs (y = 0.0082x − 0.0034;R2 = 0.1202; ANOVA notsignificant) (a) and�-blockers (y = 0.1386x − 0.1709;R2 = 0.4301;ANOVA significant;p < 0.02) (b) and octanol–water partition coeffi-cients of the compounds (logKow); calculated and measured valuesare given in different symbols. Acute toxicity ofDaphnia magnarefers to immobilization after 48 h (LC50 value).References—acutetoxicity: Calleja et al. (1993), Lilius et al. (1994), Henschel etal. (1997), Halling-Sorensen et al. (1998), Huggett et al. (2002),Brooks et al. (2003), Cleuvers (2003), Villegas-Navarro et al. (2003),Cleuvers (2004), Ferrari et al. (2004), Hernando et al. (2004),Marques et al. (2004a,b). logKow, in between parentheses: acetyl-salicylic acid (1.13) (Sanderson et al., 2003); salicylic acid (2.26)(Hansch et al., 1995); diclofenac (4.51), ibuprofen (3.97) (Avdeefet al., 1998); naproxen (3.18) (Cleuvers, 2004); paracetamol (0.49)(Henschel et al., 1997); atenolol (0.5) (Griffin et al., 1999); betaxolol(2.98) (Sanderson et al., 2003); metoprolol (2.15), propranolol (3.56)(Hardman et al., 1996); sotalol (0.24) (Hansch et al., 1995).

Comparison of toxicity is difficult in this case, sinceother beta-blockers, except metoprolol, were only ana-lyzed in D. magna (Hernando et al., 2004). Meto-prolol and verapamil caused the acceleration of theheart beat rate at low concentration, but lowered itat high concentrations inD. magna (Villegas-Navarroet al., 2003). For propranolol it seems that phyto-and zooplankton are more sensitive than fish.Ceri-odaphnia dubia [EC50 (48 h) = 0.8 mg/L;Ferrari et

al., 2004] displayed higher sensitivity thanD. magna[EC50 (48 h) = 1.6 mg/L; Huggett et al., 2002] orother zooplankton organisms. Within phytoplankton,the microorganismSynechococcus leopolensis reactedmost sensitive [EC50 (96 h) = 0.668 mg/L;Ferrari et al.,2004].

6.1.3. Blood lipid lowering agentsSimilar to beta-blockers, acute toxicity of lipid low-

ering agents is not extensively reported. Clofibrateshowed LC50 values in the range of 7.7–39.7 mg/Land can be classified as harmful to aquatic organisms.The fishGambusia holbrooki [LC50 (96 h) = 7.7 mg/L;Nunes et al., 2004] seems the most sensitive organ-ism to acute clofibrate concentrations studied so far.The known rodent peroxisome proliferator gemfibrozilinjected to rainbow trout led to significant increasesin fatty acyl-CoA oxidase (FOA) activity at dosesof 46–152 mg/kg/day (Scarano et al., 1994). Signifi-cant dose-related increases in peroxisomal FOA wereobserved after exposure of rainbow trout primary hep-atocytes to clofibric acid, and ciprofibrate, but not withgemfibrozil (Donohue et al., 1993). The in vitro activityin these fishes is weak.

6.1.4. Neuroactive compounds (antiepileptics,antidepressants)

The serotonin re-uptake inhibitor fluoxetine isapparently the most acute toxic human pharmaceu-tical reported so far with acute toxicity rangingf ,2F ngeo(ds thera

tics,c atico a areb tm rea-s utet inD asim

rom EC50 (48 h, alga) = 0.024 mg/L (Brooks et al.003) to LC50 (48 h) = 2 mg/L (Kummerer, 2004).or benthic organisms, acute toxicity is in the raf 15–43 mg/kg sediment [Chironomus tentans LC5010 days) = 15.2 mg/kg,Hyalella azteca LC50 (10ays) = 43 mg/kg;Brooks et al., 2003]. Fluoxetineeems to stronger affect phytoplankton than oquatic organisms.

Diazepam and carbamazepine, both antiepilepan be classified as potentially harmful to aqurganisms, because most of the acute toxicity datelow 100 mg/L. For both compounds it seems thaD.agna is affected more than other species, but theons for the higher susceptibility is not known. Acoxicity of carbamazepine was found at 17.2 mg/Laphnia and at 34.4 mg/L in midges, but growth w

nhibited at 12.7 mg/L inDaphnia and at 9.2 mg/L inidges (Thaker, 2005).


6.1.5. Cytostatic compounds and cancertherapeutics

Acute toxicity of methotrexate on highly prolif-erative species, namely the ciliateTetrahymena pyri-formis, indicated acute effects [EC50 (48 h) = 45 mg/L;Henschel et al., 1997]. Teratogenicity in fish embryoswas observed at even higher concentrations [EC50(48 h) = 85 mg/L;Henschel et al., 1997].

The acute toxicity data summarized inFig. 2showsthat 17% of the pharmaceuticals displayed an acutetoxicity below 100 mg/L, and for fluoxetine, all toxic-ity values were below 10 mg/L. On the other hand, 38%of the pharmaceuticals such as acetylsalicylic acid,betaxolol, sotalol, bezafibrate, gemfibrozil, bezafibrate,cimetidine and ranitidine displayed LC50 values higherthan 100 mg/L, which, according to EU-Directive93/67/EEC (Commission of the European Commu-nities, 1996), are classified as not being harmful foraquatic organisms. The other pharmaceuticals (45%)displayed a considerable variability of acute toxicityvalues, spreading over a wide range, thus making aclassification difficult.

Variability of data both within the same and betweendifferent species is obvious. Different actual exposureconcentrations (only nominal concentrations were usedin the determination of the endpoints), different sen-sitivities of used clones, different laboratory perfor-mances are among the reasons for variability withinthe same species (for example, clofibric acid toxicityin D. magna varies between 72 and 200 mg/L; the LC50( g/L(1 en9e lityo e tot pro-p quitel hend werL eu-r rugso

odeo pe-c ht toi cificm be

only one, additional ones (e.g. oxidative stress) comeinto play with particular pharmaceuticals. We evaluatedwhether the acute toxicity data of the different classesof pharmaceuticals correlate with the logKow of thecompound, as the lipophilicity determined by logKowis an important parameter for membrane toxicity. How-ever, no correlation was found between the logKow ofpharmaceuticals of a certain category or of all phar-maceuticals, and the acute toxicity either of a certainspecies, a group of organisms, or all of them. Thebest relation between measured and estimated logKowof one class of pharmaceuticals and acute toxicity inone species,D. magna, is depicted inFig. 3. Rea-sons for the variability of the data are probably basedon laboratory differences, nominal concentration dif-ferences, clone susceptibility differences, but also onthe fact that logKow may not be the best model forlipophilicity. This holds in particular for ionizable com-pounds, where the pH-dependent speciation is of sig-nificant influence (Fent and Looser, 1995; Looser et al.,1998).

In conclusion, acute toxicity to aquatic organisms isunlikely to occur at measured environmental concen-trations, as acute effects concentrations are 100–1000times higher than residues found in the aquatic envi-ronment. For example, the lowest acute effect concen-tration of fluoxetine was 20�g/L, whereas the highestestimated environmental concentration was 0.01�g/L;the lowest acute effect of salicylic acid was 37 mg/L,whereas the highest environmental concentration was∼ inc

6

overl lifec ro-p por-t andw n.T atet theq ntsa ide-l siso dif-f rely

48 h) of acetylsalicylic acid varies between168 mCalleja et al., 1994) and 1468 mg/L (Lilius et al.,994); the LC50 (24 h) of diazepam varies betwe.6 mg/L (Calleja et al., 1993) and 10000 mg/L (Callejat al., 1994)). Depending on the quantity and quaf data, ranges of acute toxicity values span on

wo orders of magnitude, in some cases such asranolol or diazepam, the species differences are

arge, spanning three to four orders of magnitude. Wifferent categories are compared, a tendency of loC50 (EC50) values is found for beta-blockers and noactive drugs as compared to antiinflammatory dr various other compounds.

Often, acute toxicity is related to non-specific mf actions, and not to mechanisms involving sific target molecules. The compounds are thougnteract with cellular membranes leading to unspe

embrane toxicity. This general mechanism may

60�g/L. Therefore, acute toxicity is only relevantase of spills.

.2. Chronic effects

Many aquatic species are continuously exposedong periods of time or even over their entireycle. Evaluation of the chronic potential of micollutants, e.g. pharmaceuticals, is therefore im

ant. However, there is a lack of chronic data,here available, chronic toxicity is marginally knowhe available chronic data do often not investig

he important key targets, nor do they addressuestion in different organisms. Toxicity experimere usually performed according to established gu

ines. More specific investigations including analyf possible targets of the pharmaceutical, or over

erent life stages, are lacking, or have only ra


been performed. Moreover, life-cycle analyses are notreported, except for EE2 (Lange et al., 2001; Parrottand Blunt, 2005), and toxicity to benthic and soil organ-isms have very rarely been evaluated. In this chapter,we review the current literature according to the differ-ent pharmaceutical classes and summarize the data inFig. 4.

The best knowledge exists for thesynthetic steroidEE2 contained in contraceptive pills, showing estro-genic effects at extremely low and environmentallyrelevant concentrations. This steroid has been shown

Ftcft(((

in many fish to induce estrogenic effects at extremelylow concentrations. Exposure of fathead minnowsover their life cycle indicates reproductive effects atlow concentrations of EE2 (Lange et al., 2001). TheNOEC values of the F0 generation F1 embryo hatch-ing success and larval survival were≥1 ng/L. Malefish exposed to EE2 at 4 ng/L failed to develop nor-mal secondary sexual characteristics and the sex ratiowas altered. No testicular tissue was observed in anyfish exposed to EE2 at 4 ng/L. A recent study showsvitellogenin (VTG) induction in fathead minnows withan EC50 value as low as 1 ng/L; EE2 was 25–30 timesmore potent than estradiol (Brian et al., 2005), con-firming previous reports on VTG induction at con-centrations between 0.1 and 1 ng/L (Pawlowski etal., 2004). Decreased egg fertilization and sex ratio(skewed toward females), both of which were sig-nificantly affected at extremely low concentrations of0.32 ng/L EE2 (Parrott and Blunt, 2005). The next mostsensitive parameter was demasculinization (decreasedmale secondary sex characteristic index) of malesexposed to an EE2 concentration of 0.96 ng/L. Fulllife-cycle exposure of zebrafish to 3 ng/L EE2 leadto elevation of VTG and caused gonadal feminiza-tion in all exposed fish and thus inhibited reproduction(Fenske et al., 2005). Life-long exposure of zebrafishto 5 ng/L in the F1 generation caused a 56% reduc-tion in fecundity and complete population failure withno fertilization. Infertility in the F1 generation wasdue to disturbed sexual differentiation with males hav-ie

eena ua-t ult,b of ag in-g areo osc rd f

ig. 4. Chronic toxicity of 10 different pharmaceuticals, belongingo different therapeutic classes. Given are lowest observed effect con-entrations (LOEC) and no observed effect concentrations (NOEC)or different aquatic organism, different endpoints and exposureimes. See text for details.References: Webb (2001), Huggett et al.2002), Brooks et al. (2003), Ferrari et al. (2003, 2004), Cleuvers2004), Henry et al. (2004), Kummerer (2004), Marques et al.2004a,b), Schwaiger et al. (2004)andTriebskorn et al. (2004).

m a-t d ont ccu-r s inT bed hisi

ng no functional testes and intersex gonads (Nasht al., 2004).

In hazard and risk assessment, the ratio betwcute to chronic toxicity is often taken for eval

ion of chemicals. For pharmaceuticals, this is difficecause only very rarely, a systematic analysisiven drug in both acute and chronic toxicity in a sle species is performed. Apart from EE2, therenly a few NSAID, from which acute to chronic ratian be deduced.Table 3shows that even for similarugs, these ratios inD. magna vary by two orders oagnitude. For all other drugs, only partial inform

ion is available on a given species. Ratios derivehe basis of a number of different species are not aate, giving questionable information. The exampleable 3confirm again that chronic toxicity cannoterived from acute toxicity by simple calculations. T

s often neglected in risk assessment.


Table 3Ratio between acute and chronic toxicity inDaphnia magna andCeriodaphnia dubia (48 h/21days)

Drug Acute (mg/L) Chronic (mg/L) Ratio

Acetylsalicylic acid 1293.1 1.4 924Salicylic acid 1031.7 13.3 77Clofibrate 28.2 0.01 2820Naproxen 66.4 0.33 201Naproxen Na 43.6 0.68 64

Data afterMarques et al. (2004a,b)(acetylsalicylic acid and salicylicacid,D. magna), Webb (2001)(clofibrate,D. magna) andIsidori etal. (2005)(naproxen and naproxen Na,Ceriodaphnia dubia).

6.2.1. Analgesics and non-steroidalantiinflammatory drugs

NSAID inhibit the synthesis and release ofprostaglandins via COX inhibition and these com-pounds are the most consumed category of drugs.About NSAID commonly found in the aquatic envi-ronment, most chronic data are reported. Acetylsali-cylic acid affected reproduction inD. magna andD.longispina at concentrations of 1.8 mg/L (Marques etal., 2004a). Diclofenac is commonly found in wastew-ater at median concentration of 0.81�g/L (Ternes,1998) whereas the maximal concentration in wastewa-ter and surface water is up to 2�g/L (Stumpf et al.,1996; Ternes, 1998; Schwaiger et al., 2004). Tradi-tional chronic toxicity studies with diclofenac werereported in invertebrates (Ferrari et al., 2003, 2004).A recent study demonstrated chronic histopathologi-cal effects in rainbow trout after 28 days of exposure.At the LOEC of 5�g/L renal lesions (degeneration oftubular epithelia, interstitial nephritis) and alterationsof the gills occurred in rainbow trout (Schwaiger etal., 2004), and subtle subcellular effects even at 1�g/L(Triebskorn et al., 2004). Impairment of renal and gillfunction is likely to occur after long-term exposure.The kidney was also found to be a target of diclofenacin vultures, acute renal failure was probably the rea-son for the visceral gout (Oaks et al., 2004) and theoccurrence of extensive deposits of uric acid on andwithin internal organs (Gilbert et al., 2002). In zebrafishembryos, no effect of diclofenac on embryonic devel-o 1 and2 fd liverwe hee

6.2.2. Beta-blockersAs fish contain�2-receptors in heart and liver

(Gamprel et al., 1994) and probably in reproductivetissues (Haider and Baqri, 2000), unspecific antago-nists such as propranolol may be active in fish. Infact, propranolol indicated chronic toxicity not onlyon the cardiovascular system, but also on reproduc-tion. The no-observed-effect-concentration (NOEC)and lowest-observed-effect-concentration (LOEC) ofpropranolol affecting reproduction inC. dubia were125 and 250�g/L, and reproduction was affected after27 days of exposure inH. azteca at 100�g/L (Huggettet al., 2002). In fish O. latipes, significant changes inplasma steroid levels occurred after 14 days of expo-sure. The number of eggs released by fish was reducedat 0.5�g/L after a 4-week exposure to 0.5 and 1�g/L,but not at 50 and 100�g/L (Huggett et al., 2002).No alteration in vitellogenin levels was observed. Itwas suggested that alteration in sex steroids let todecreased oxytocin excretion, which could decreasethe number of eggs released. Propranolol was also ana-lyzed in invertebrates. LOEC and NOEC for differentorganisms span several orders of magnitude (Fig. 4),partly due to differences between laboratories, but alsospecies differences. These data should be compared tothe environmental concentrations; propranolol, meto-prolol and nadolol were identified in U.S. wastewa-ter samples up to 1.9, 1.2 and 0.36�g/L, respectively(Huggett et al., 2002).

6ates

h Thefdcl /L]( dp 113-t tionf d byo wella cripti

6rba-

m itors

pment was observed, except delayed hatching atmg/L (Hallare et al., 2004). Additional side effects oiclofenac have been observed in humans in theith degenerative and inflammatory alterations (Bankst al., 1995), in lower gastrointestinal tract and in tsophagus (Bjorkman, 1998), but not in fish.

.2.3. Blood lipid lowering agentsData on this class of compounds are rare. Fibr

ave been evaluated by traditional toxicity tests.ollowing NOEC were found for clofibric acid inC.ubia [NOEC (7 days) = 640�g/L], the rotiferB. caly-iflorus [NOEC (2 days) = 246�g/L], and in earlyife stages of zebrafish [NOEC (10 days) = 70 mgFerrari et al., 2003). Gemfibrozil occurred in bloolasma of goldfish after exposure over 14 days at

imes higher levels than in water (bioconcentraactor of 113). Plasma testosterone was reducever 50% after exposure to 1.5 and 10 mg/L, ass levels of steroid acute regulatory protein trans

n goldfish testes (Mimeault et al., 2005).

.2.4. Neuroactive compoundsMost data were reported for the antiepileptic ca

azepine and selective serotonin re-uptake inhib


(SSRI), other neuroactive compounds were very rarelyor not evaluated (Fig. 4). Traditional toxicity testsshowed chronic toxicity of carbamazepine inC. dubia[NOEC (7 days) = 25�g/L], in the rotifer B. caly-ciflorus [NOEC (2 days) = 377�g/L], and in earlylife stages of zebrafish [NOEC (10 days) = 25 mg/L](Ferrari et al., 2003). Carbamazepine is considered car-cinogenic in rats but is not mutagenic in mammaliancells. Sublethal effects occurred inDaphnia at 92�g/Land the lethal concentration in zebra fish was 43�g/L(Thaker, 2005). In a study with the cnidarianHydravulgaris, diazepam was shown to inhibit polyp regen-eration at 10�g/L (Pascoe et al., 2003).

Most chronic studies focused on SSRI. Serotonin isa neurotransmitter found in lower vertebrates and inver-tebrates, and SSRI may adversely influence the func-tion of the nervous and associated hormonal systemsof these organisms as well. Besides having importantfunctions as a neurotransmitter, serotonin may directlyact on the immune system, alters appetite, influencesbehavior and modulates sexual function. The roleof serotonin in reproduction varies between differentphyla and effects of SSRI as well.Fong (1998)foundthat SSRI (fluvoxamine, paroxetine) led to induction(at 10 nM to 100�M) and fluoxetine to potentiation (at5�M, and if co-applied with 7–100�M serotonin, butnot at other concentrations) of parturition in fingernailclams.Fong (1998)found an induction of spawning inzebra mussels by fluvoxamine concentrations as low as0.032�g/L. Induction of mussel spawning point to ani tes,s andj en-es3dru x-e ectso -b dayso line,te ageaer

In medaka (O. latipes), serotonin induced oocytematuration (Iwamatsu et al., 1993), but a contraryaction was reported in mummichog (F. heteroclitus)(Cerda et al., 1998). Serotonin was indicated to poten-tiate effects of gonadotropin-releasing hormone ongonadotropin release from the pituitary (Khan andThomas, 1994). When medaka were exposed for 4weeks to fluoxetine concentrations of 0.1–5�g/L,vitellogenin plasma content, plasma steroids, fecun-dity, egg fertilization or hatching rate were not affected(Foran et al., 2004). This indicates no reproductionimpairment in this fish up to 5�g/L fluoxetine. Takentogether the chronic effects of SSRI on reproductionof fish and invertebrates are not yet clear, interferencewith reproduction occurred at much higher concentra-tions than measured in surface waters.

Chronic data on various other compounds are lack-ing, although they have been shown to occur in consid-erable amounts in surface waters (Fig. 2). This holdsin particular for fish. For the anticancer compound,tamoxifen, chronic data are found forAcartia tonsa[EC50 = 49�g/L; Andersen et al., 2001]. Various mor-phological and developmental effects (early embryonicmortality) were induced in sea urchin embryos afterexposure to 10−8 to 10−5 M tamoxifen, which cor-responded to oxidative stress. ROS production wasincreased and lead to oxidative damage and it is thoughtto represent a pro-oxidant mode of action explainingcarcinogenicity in humans and rodents (Pagano et al.,2001).

aro-m fore thee ntalc headm tlyr dro-z edt andi sisi La bryoh istol-of andm ecro-s ronei d an

nterference with serotonin action, as in invertebraerotonin may stimulate ecdysteroids, ectysoneuvenile hormone, responsible for controlling oogsis and vitellogenesis (Nation, 2002). A reproductivetimulation was also found inD. magna exposed to6�g/L fluoxetine for 30 days, and inC. dubia fecun-ity was increased at 56�g/L (Flaherty et al., 2001), buteduced in another study (Brooks et al., 2003). An eval-ation of five SSRI (fluoxetine, fluvoxamine, parotine, citalopram, sertraline) showed negative effn C. dubia reproduction by reduction of the numer of neonates or brood per female after 7–8f exposure. For the most active compound, sertra

he LOEC was 45�g/L and the NOEC 9�g/L (Henryt al., 2004). Fluoxetine has been detected in sewnd stream water at concentrations of 12 ng/L (Kolpint al., 2002) and 99 ng/L (Metcalfe et al., 2003b),espectively.

The antiandrogenic compound flutamide andatase inhibitor fadrozole were also analyzedffects in fish, mainly as a positive control forvaluation of effects suspected for other environmehemicals. Short-term reproduction assays in fatinnows show that flutamide at 0.9 mg/L significan

educed male sex characteristics in male fish. Faole significantly inhibited ovarian growth and inducestis growth at 0.05 and 0.96 mg/L after 21 days,nhibited VTG in females and induced VTG synthen males (Panter et al., 2004). Flutamide at 0.5 mg/lso reduced fecundity of the fish after 21 days. Ematch was reduced and alterations in gonadal hgy were observed (Jensen et al., 2004). Ovaries from

emales indicated a decrease in mature oocytesales exhibited spermatocyte degeneration and n

is. Concentration-dependent VTG and testostencrease were observed in females. Flutamide ha


antiandrogenic effect and reduced fecundity, yet atrather high concentrations. Moreover, in adult maleguppy, reduction in ejaculated sperms, reduced sex col-oration and smaller testes occurred. The male courtshipbehavior was also disrupted at 1 and 10 mg/kg in feed(Baatrup and Junge, 2001). The aromatase inhibitorfadrozole reduced fecundity after 21 days at waterconcentrations of 10 and 50�g/L and inhibited brainaromatase activity (Ankley et al., 2002). In females aconcentration dependent reduction in plasma estradioland VTG was observed. In males, androgens in plasmawere significantly increased and resulted in a markedaccumulation of sperm in the testes.

6.3. In vitro studies

Several pharmaceuticals have been investigated inin vitro systems. They were mainly analyzed for acutecytotoxicity in fish cell lines and in primary fish cellcultures. Cytotoxicity of clofibrate, fenofibrate, carba-mazepine, fluoxetine, diclofenac, and propranolol tothe fish cell line PLHC-1 (hepatoma cells derived fromtopminnow) and primary cultures of trout hepatocyteswas reported (Laville et al., 2004). Fibrates are knownto enhance�-oxidation of lipids, which increases theamount of reactive oxidative species (ROS) in cells.Fenofibrate [EC50 (24 h) = 3.25 mg/L] and clofibrate[EC50 (24 h) = 0.46 mg/L] were the most active com-pounds (Laville et al., 2004). Cytotoxicity was higher inPLHC-1 than in primary hepatocytes. Oxidative stressi atli ]w idesc calsw meP canb ngt istp ryh inhi-b

ec ica ob-l t off itroc d-

ies with highly proliferating ciliates. The particularmode of action of methotrexate [EC50 (48 h) = 3 mg/L;Henschel et al., 1997] may negatively interact with cellproliferation and therefore survival. Sensitivity of cellsto toxicants may vary within species, as demonstratedby a direct comparison between cytotoxicity on fish andrat cell lines (Rau et al., 2004), or depending on theirorigin, e.g. PLHC-1 cell lines are more sensitive thantrout primary hepatocytes (Laville et al., 2004). Someof the differences may be based on the difference inthe ability of the cells to metabolize toxicants. These invitro studies indicate their usefulness for the acute toxi-city evaluation, but also for investigations of the modesof action of pharmaceuticals including chronic toxicityparameters. Among the advantages of in vitro systemsbased on fish cells or reporter gene systems are theirpotential for screening and first evaluation of potentialtoxicity (Fent, 2001). They are important alternativesto animal testing able to identify general toxicity andspecific cellular targets and processes, and they areeconomic.

6.4. Toxicity of pharmaceutical mixtures andcommunity effects

There are only a few studies dealing with the effectsof mixtures of pharmaceuticals.Cleuvers (2003, 2004)has evaluated the ecological potential of antiinflam-matory drugs and of diverse acting pharmaceuticalsin different sets of biotests using different aquatico o-f atedu em sin-g hem iona achc

ew ined id,c olol,d m-b andc tiona ts,t hen

s thought to be responsible for the cytotoxicity,east for these fibrates (Laville et al., 2004). Cytotox-city of fluoxetine [EC50 (24 h, PLHC-1) = 1.73 mg/Las also mediated in part by oxidative stress. Besytotoxicity and ROS production, the pharmaceutiere analyzed for their potential to induce cytochro4501A monooxygenase activity (CYP1A), whiche regarded as important for chronic toxicity. Amo

he tested drugs, the�-adrenergic receptor antagonropranolol was the only CYP1A inducer in primaepatocytes, the other six pharmaceuticals lead toition of basal activity (Laville et al., 2004).

Furthermore,Henschel et al. (1997)evaluated thytotoxicity of salicylic acid, paracetamol, clofibrincid and methotrexate to BF-2 fish cell line (fibr

asts derived from bluegill sunfish). For three ouour compounds the concentrations inducing in vytotoxicity were lower as compared to in vivo stu

rganisms. A mixture of NSAID (diclofenac, ibupren, naproxen, acetylsalicylic acid) has been evalusing acuteDaphnia and algal tests. Toxicity of thixture was found at concentrations at which thele compound showed no or only little effects. Tixture toxicity followed the concept of concentratddition, which means that the concentrations of eompound behaved in an additive fashion.

Acute toxicity tests usingD. magna, alga (Desmod-smus subspicatus) and macrophyte (Lemna minor)ere performed to analyze for acute toxicity of nrugs having different modes of action (clofibric acarbamazepine, ibuprofen, propranolol, metopriclofenac, naproxen, captopril, metformin). The coined effects of two substances, clofibric acidarbamazepine, followed the concept of concentraddition in theDaphnia test, whereas in the algal tes

he concept of independent action was adequate. W


a combination of NSAID, ibuprofen and diclofenac,was analyzed, the effect on algae followed the con-centration addition concept, whereas forDaphnia, thecombination effect was stronger. These data indicatethat for the acute toxicity of these pharmaceuticals,concentration addition can be assumed, which meansthat the concentration of each individual pharmaceuti-cal has to be added for the combination effects. Thisimplies that compounds occurring at concentrationsbelow their individual NOEC can nevertheless con-tribute to the total effect of the mixture.

Only few pharmaceuticals have been analyzed inecologically more realisticmodel ecosystems, namely,microcosms and mesocosms. In two recent studies,outdoor aquatic microcosms of a total volume of12,000 L containing water and sediment were usedto analyze the effects of combination of pharmaceu-ticals. (Brain et al., 2004a) evaluated the effects ofcombinations of eight pharmaceuticals at three con-centration levels on macrophytesLemnea gibba andMyriophyllum sibiricum over a 35 days period. Ator-vastatin, a blood lipid regulator, was among antibi-otics the pharmaceutical eliciting phytotoxicity. Usingsimilar microcosms effects on phyto- and zooplank-ton were assessed after exposure for 35 days at threeconcentrations to two pharmaceuticals (ibuprofen, flu-oxetine) and an antibiotic (ciprofloxacin) (Richards etal., 2004). The microcosms contained periphyton, phy-toplankton, zooplankton, algae and benthic communi-ties, and in addition, juvenile sunfish were exposed inm hyto-p ium( entl at-m utn hal-i nts,l sedg andz anceo icali ind therc eu-t n-c cindr )

concluded that a low probability exists that these threepharmaceuticals are currently present in surface watersat concentrations negatively affecting aquatic commu-nities. By comparing calculated whole-body therapeu-tic doses – and not human and fish plasma levels – theauthors note that all responses occurred at levels wellbelow the equivalent pharmacologically active concen-trations in mammals. Concentrations of pharmaceuti-cals in fish can reach significantly higher concentra-tions in plasma than in the ambient water (Mimeault etal., 2005).

7. Comparison of environmentalconcentrations and ecotoxicological effectsconcentrations

The potential risk of a substance to the environmentis often characterized by comparing the Predicted Envi-ronmental Concentration (PEC) with the Predicted NoEffect Concentration (PNEC). PEC of pharmaceuticalsare often estimated using calculations, which includeusage or sales figures, population density, wastewaterproduction and dilution in watersheds to generate likelyconcentrations in surface waters (Halling-Sorensen etal., 1998; Jones et al., 2002; Straub, 2002; Sandersonet al., 2003; Bound and Voulvoulis, 2004). Due tothe lack of experimental data (in particular chronic)in the public domain on the ecotoxicity of pharma-ceuticals, estimation of PNEC, and therefore hazarda iblet forl able,a aveb gicalt 002;S n-t tives ns,f am(b s orp uffi-c entso

ntlya ax-i EC

esh cages. Species abundance and number of plankton and zooplankton were affected at the med60–100�g/L each compound), and high treatmevel (600–1000�g/L each), whereas at the low tre

ent (6–10�g/L each), only trends were visible, bo significant effects occurred. Unexpected high let

ty occurred in fish at the high and medium treatmeethality was observed in plants in addition to decrearowth. Decreased diversity of both phytoplanktonooplankton communities and increased abundf both communities may have important ecolog

mplications. However, the cause of the declineiversity and the other effects was unclear (wheaused directly or indirectly and by what pharmacical having different modes of action). Maximal coentrations of ibuprofen, fluoxetine and ciprofloxaetected in the U.S. were 1.0, 0.012 and 0.03�g/L,espectively (Kolpin et al., 2002).Richards et al. (2004

nd risk assessment, is difficult or even imposso perform. In the open literature or databases,ess than 1% of pharmaceuticals data are availnd only a small number of new pharmaceuticals heen undergone risk assessment using ecotoxicolo

ests (Halling-Sorensen et al., 1998; Jones et al., 2anderson et al., 2003) In the absence of experime

al data, information is often derived from quantitatructure–activity relationships (QSAR) predictioor example by applying the EPA’s ECOSAR progrJones et al., 2002; Sanderson et al., 2004a). Whileeing a pragmatic approach for identifying hazardrioritizing critical substances, this concept is not siently precise for accurate hazard and risk assessmf pharmaceuticals.

Here, we summarize and compare the currevailable empirical data in the open literature on m

mal STP effluent concentrations with chronic LO


Fig. 5. Comparison between maximal concentrations of pharmaceuticals in treated wastewater and their chronic toxicity in aquatic organisms.(a) Lowest observed effect concentrations (LOEC); (b) no observed effect concentrations (NOEC) for different aquatic organism, differentendpoints and exposure times. References seeFig. 1(wastewater concentrations) andFig. 4(chronic toxicity).

and NOEC concentrations of individual pharmaceuti-cals (Fig. 5). This approach is based on experimentaldata allowing to prioritize pharmaceuticals accordingto their ecotoxicological potential and to gain knowl-edge about the worst case situation. As can be deducedfrom Fig. 5, LOEC and NOEC values of the pharma-ceuticals for different aquatic organisms are about oneto two orders and two orders of magnitude, respec-tively, higher than maximal concentrations in STPeffluents. For diclofenac, the LOEC for fish toxic-ity was in the range of wastewater concentrations,whereas the LOEC of propranolol and fluoxetine forzooplankton and benthic organisms were near to maxi-mal measured STP effluent concentrations. This showsthat for diclofenac, propranolol and fluoxetine the mar-gin of safety is narrow, and chronic effects at highlycontaminated sites cannot be completely ruled out, inparticular, when the combined effects of pharmaceuti-cal mixtures are taken into account. However, mediansewage effluent concentrations are lower and dilution inreceiving waters result in lower levels in surface watersreducing the environmental risk. It should be noted,however, that more experimental data on chronic toxi-city and on the bioaccumulation potential is needed tofully judge the environmental risk posed by individualpharmaceuticals.

8. Discussion

Pharmaceuticals have been tested in traditionalways. A set of mainly acute toxicity tests using tradi-tional species such as an algae (mainlyScenedesmusquadricauda), zooplankton (D. magna) and fish(species according to OECD guidelines) has been per-formed. In general, only very few pharmaceuticals havebeen assessed for acute and chronic toxicity in fish.Moreover, only a few pharmaceuticals have been ana-lyzed for chronic toxicity, again in the traditional wayaccording to guidelines (OECD, U.S. EPA). Based onthese studies, no one would probably have been ableto anticipate the current population decline of threespecies of vultures due to diclofenac exposure. Further-more, these tests alone are not sufficient for deriving anaccurate profile of the possible hazards and risks of thepharmaceutical in question. Current tests cover only asmall set of laboratory organisms, which are often notsensitive enough and often not able to unravel adverseeffects of pharmaceuticals. As a consequence, morespecific tests are needed. Only chronic toxicity inves-tigations using more specific toxicity parameters willlead to a more meaningful ecological risk assessment.The following working hypotheses should be addressedin future ecotoxicological investigations:


1. Pharmaceuticals as biologically active compoundsmay have similar (chronic) effects in non-mammalian animals (and even plants) as in mam-mals as target receptors and/or biomolecules aresimilar and conserved during evolution. Therefore,similar adverse (chronic) effects as in humans andmammals may occur in lower vertebrates and inver-tebrates.

2. Some pharmaceuticals may have unexpected(chronic) effects in lower organisms due to biologi-cal differences in pharmacodynamics, pharmacoki-netics and physiology.

3. In vitro studies of pharmaceuticals are importantfor screening, elucidating the modes of action innon-target organisms, and designing specific in vivostudies.

One approach to address these hypotheses is toinclude histopathological investigations in chronic fishtoxicity studies. By focusing on specific tissues andorgans, more detailed answers about possible adverseeffects may be obtained. This is exemplified by a studyon chronic effects of diclofenac in fish (Schwaiger etal., 2004; Triebskorn et al., 2004). Another approachis to use the existing knowledge about possible sideeffects of the compound of interest in mammalsand humans for the design of specific analysis inaquatic organisms. Furthermore, known drug–druginteractions in humans may be relevant for compoundmixtures in the environment. Both are based on theh y bei ors,b rvedi ears smson tors( chaa ndg ratesam on,a omeP isms( siso tedt

• specific and identical targets (biomolecules, tissues,organs): target specificity

• known adverse side effects in humans and mammals:side effect specificity

• general chronic effects for accounting physiologicaldifferences: species specificity

Our proposed strategy for future research on theecotoxicology of pharmaceuticals is exemplified by afew examples. When the ecotoxicity of NSAID is stud-ied, effects on inhibition of prostaglandin synthesis andCOX inhibition should be addressed in lower organ-isms, and at the same time, on side effects alreadyknown in mammals. Diclofenac has been known forcausing side effects on the kidney (and other organssuch as liver) in mammals, subsequently being foundin vultures (Oaks et al., 2004), and fish (Schwaiger etal., 2004). Cardiovascular pharmaceuticals should beanalyzed for their possible effects on the cardiovascu-lar system in lower vertebrates. Lipid lowering agentssuch as fibrates are also known to act by enhancing orreducing PPAR (Kliewer et al., 1997). These nuclearreceptors play key roles in the catabolism and storage offatty acids and are important for blood lipid regulation.Indeed, PPAR’s are affected in amphibians (Klieweret al., 1997) and fish by clofibrate, benzafibrate andfenofibrate (Ruyter et al., 1997).

Beta-blockers bind to the beta-adrenergic recep-tors and block its activation by physiological agonists.These receptors are located in mammals in many tis-s reasei r ins mea(r tb afterc ,2o cedra

eirt ase(W wera ions.S lant

ypothesis that targets of the pharmaceutical madentical or similar in lower organisms as receptiochemical pathways and enzymes are conse

n evolutionary terms. This holds true for nuclteroid receptors that are very similar in organif different evolutionary levels (Wilson et al., 2004),uclear peroxisome proliferator-activated recepPPAR’s) (Escriva et al., 1997), adrenoceptors sus�1- and�2-receptors (Nickerson et al., 2001), butlso for insulin receptor, insulin-like growth factor alucagon receptors being present in lower vertebnd invertebrates (Navarro et al., 1999). Also, basicechanisms like signal transduction, cell divisind key metabolizing enzymes such as cytochr450s are conserved in a large variety of organ

Nelson et al., 1996). As a consequence, analyf pharmaceuticals should specifically be direc

o

ues including heart, and its blockade causes a decn heart rate and contraction. Beta-blockers diffepecificity to the different receptor subtypes, sore non-specifically acting on�1- and �2-receptorse.g. propranolol), while others are specific for the�1-eceptor subtype (e.g. atenolol). InD. magna, heareat rate, fecundity and biomass were reducedhronic exposure to 0.11 mg/L (Dzialowski et al.003), although it is not know whether�2-receptorsccur. Long-term exposure to propranolol redueproduction inC. dubia at 250�g/L and inH. aztecat 100�g/L (Huggett et al., 2002).

A class of antihyperlipidemic drugs inhibit tharget enzyme hydroxymethylglutaryl-CoA reductHMG-CoA reductase) in mammals (Seiler, 2002).

hether these enzymes are also inhibited in lonimals should be addressed in future investigaturprisingly, atorvastatin was even found in a p


(duckweed,Lemna gibba) to have effects, but the modeof action is unclear (Brain et al., 2004b).

Many antineoplastic drugs used in cancer therapyhave a high mutagenic and cancerogenic potential.Parent compounds are often bioactivated leading toformation of mutagenic metabolites (e.g. cyclophos-phamide, ifosfamide). In case organisms in the envi-ronment are able to metabolize these pharmaceuticals,enhanced mutation frequencies and cancer risk willresult. In addition, these drugs often have significantside effects on humans such as nausea, cytotoxicity,reduction in proliferation of cells in various tissues etc.One would expect mutagenicity and cancerogenicity tooccur in exposed aquatic organisms as well. However,such studies are lacking besides the analysis of hospitalwastewater, in which the genotoxic potential was basedon antibiotics such as ciprofloxacin (Hartmann et al.,1998).

It should be noted, however, that besides known tar-gets additional or other target tissues and organs may beaffected alternatively. This would result inunexpectedeffects not targeted by the investigations. Examplesare effects on sex hormones in blood plasma of fishand reduced reproduction inC. dubia and H. aztecainduced by the beta-blocker propranolol after long-term exposure (Huggett et al., 2002), and the effects ofserotonin-re-uptake inhibitors on reproduction of mol-lusks (Fong, 1998; Fong et al., 1998). As the effectsof the antiestrogen tamoxifen indicate, pharmaceuti-cals may have not only one, but multiple modes ofa aseo -p cts.H nsesw nict ro-d nsivea eu-t thee oft n-m lysess fulla ratedd cess( hav-i ts inm

In vitro studies are important for screening andevaluation of possible cellular targets in ecotoxicology(Fent, 2001). They are also important in the reductionof animal experiments, in conjunction with other pro-posed new strategies (Hutchinson et al., 2003). Effectsof pharmaceuticals have been evaluated in fish primarycells and fish cell lines indicating this potential (Lavilleet al., 2004). We assume that investigating pharmaceu-ticals in in vitro systems will not only allow a reductionof animal experiments, but also a better and more accu-rate characterization of possible targets of pharmaceu-ticals. These test systems not only allow the analysisof specific receptor interactions and target enzymes inanimal and plant cells, but also a rapid screening of alarge number of compounds.

Pharmaceuticals are analyzed for possible ecotoxi-cological effects as single compounds and only rarelyasmixtures (Cleuvers, 2003). However, as other envi-ronmental pollutants pharmaceuticals are present inthe environment in mixtures. Effects of mixtures mostprobably follow the concept of concentration addition,hence, the overall toxicity is the result of the sum of theindividual concentration of each compound. Thereforeeffects may occur even at the NOEC of individual com-pounds. It should also be recognized that even subtlechanges of normal homeostasis including behavioralalterations may have direct and indirect effects, evenif only minor ones, that eventually result in significantdeteriorating effects on a species or population in theecological context. The extreme case of the dramaticp resi tiond anceo the1

9

ma-c er ist reaset rings ovalt iza-t canb t thei ble,

ction, such as oxidative damage in addition in cf tamoxifen (Pagano et al., 2001). This fact comlicates the strategy to analyze for chronic effeowever, many of these unexpected chronic respoill be elucidated in the context of careful chro

oxicity analyses including histopathology and repuction. However, such analyses are more expend probably only justified for important pharmac

icals occurring in significant concentrations innvironment. But in the light of the limitations

raditional (acute) toxicity testing for use in enviroental risk assessment, more specific toxicity ana

hould be performed in forthcoming studies, takingdvantage of the available knowledge that is geneuring the pharmaceutical drug development proe.g. mechanisms of action, pharmacokinetic beor and metabolism, target organs and side effec

ammals).

oisoning and population declines of Indian vultus a case in point. The dimension of this populaecline has no parallel in birds since the disappearf peregrine falcons and other predatory birds in960s due to the pesticide DDT.

. Conclusions and future directions

One important aspect to solve the load of phareutical residues in wastewater and surface wato optimize STP processes. There is a need to inche knowledge about the fate of pharmaceuticals duewage treatment for implementation of better remechniques. Future work on STP treatment optimion will show to what extend pharmaceuticalse removed from wastewater and to what exten

mplementation of an improved technology is feasi


taking into account other macro- and micro-pollutantsas well as the broad variety of complex wastewatermatrices.

Our present knowledge about residues of pharma-ceuticals in aquatic systems indicate that they areunlikely to pose a risk for acute toxicity. Environ-mental concentrations are in the range of 103 to 107

times lower than known LC50 or EC50 values (ratioof 103 between lowest acute effect of fluoxetine andhighest environmental concentration; difference of 107

between highest LC50 of diazepam and highest envi-ronmental concentration). However, as the collapse ofvulture populations in the Indian subcontinent indi-cates, important adverse effects can occur under certaincircumstances.

There is a general lack of chronic toxicity dataon pharmaceuticals, in particular in fish. Many phar-maceuticals need more investigation about potentiallong-term ecotoxicological effects, particularly withrespect to potential disturbances in hormonal home-ostasis (endocrine disruption), immunological status,or gene activation and silencing during long-termexposure. For better understanding of possible effects,a mechanism-based approach focused on targetmolecules, tissues and organs should yield moremeaningful results and insights than traditional acutetoxicity testing. Current data on acute and chronictoxicity of pharmaceuticals support to the conclusionthat more target- or biomolecule-oriented, or mode-of-action-based investigations, will allow more relevanti ro-d ing.O on-m fects.I othe andf ar-m orei ualp ionsc thea

on-m thed am-i ffer-e andh ioral

changes, to name some key targets, may have far reach-ing effects on the population level. This has becomeevident for endocrine disrupters such as steroid hor-mones used in contraceptives resulting in importantadverse effects at environmentally relevant concentra-tions (Jobling et al., 1998; Lange et al., 2001; Thorpeet al., 2003; Parrott and Blunt, 2005).

Comparison of available chronic toxicity data withenvironmental concentrations indicate that for mostinvestigated pharmaceuticals concentrations are toolow in aquatic systems to induce chronic effects on tra-ditional laboratory organisms such as inhibition of algalgrowth and reproduction inDaphnia. For diclofenac,the LOEC for fish toxicity on an organ level wasin the range of wastewater concentrations, however(Schwaiger et al., 2004), whereas the LOEC of pro-pranolol and fluoxetine for zooplankton and benthicorganisms were near to maximal measured STP efflu-ent concentrations. Whether or not the margin of safetyis narrow for additional human pharmaceuticals shouldbe investigated in future studies. The future require-ment of chronic testing with algae, daphnids and fishinstead of only traditional acute toxicity studies is animportant step forward (EMEA, 2005). Moreover, thepotential of combined effects of pharmaceutical mix-tures should be addressed. In the ecological context,subtle changes and disturbances may have negativeconsequences for the organism’s fitness. As a conse-quence much more should be known about the potentialfor chronic effects of pharmaceuticals in the aquatics

A

dTu rn,S orn,N -LaR portbF g-b atlya ted,S thea ont

nsights into effects on survival, growth and repuction than traditional standard ecotoxicity testften, similar target biomolecules are present in nammalian organisms and so are the adverse ef

n vitro systems are very important tools for blucidating modes of action in lower vertebrates,

or screening of the ecotoxicological potential of phaceuticals prior to fish toxicity testing. Unless m

s known about possible chronic effects of individharmaceuticals and mixtures thereof, conclusoncerning hazards or risks of pharmaceuticals toquatic ecosystem are premature.

Drugs may also induce unexpected effects in nammalian organisms, however. This is based onifference in pharmacokinetics and pharmacodyn

cs, important parameters for occurring species dinces. Disturbances of the reproductive systemormone system, immune depression, neurobehav

ystem.

cknowledgements

We thank the Bundesamt fur Berufsbildung unechnologie (BBT), Kommission fur Technologiend Innovation (KTI-Project 7114.2 LSPP-LS), Bepringborn Smithers Laboratories (Europe) AG, Hovartis International AG, Basel, and F. Hoffmannoche Ltd., Basel, for funding this study. The supy K. Eigenmann, Novartis International AG, H. Kunzi,. Hoffmann-La Roche Ltd, and H. Galicia, Sprinorn Smithers Laboratories (Europe) AG is grecknowledged. We thank IMS Health Incorporawitzerland, for data on drug consumptions, andnonymous reviewers for constructive comments

he manuscript.


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(1) Eco Toxicology of Human Pharmaceuticals (FENT - 2006)

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