Ecotoxicological effects at contaminated sites

Toxicology 205 (2004) 223–240

Ecotoxicological effects at contaminated sites

Karl Fenta,b

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

Available online 27 August 2004

Abstract

Contamination sites pose significant environmental hazards for terrestrial and aquatic ecosystems. They are important sourcesof pollution and may result in ecotoxicological effects on terrestrial, groundwater and aquatic ecosystems. At severely contam-inated sites, acute effects occur, but the core problem lies in long-term chronic effects. Ecotoxicological effects occur at alllevels of biological organization, from the molecular to the ecosystem level. Not only certain organisms may be affected, butthe ecosystems as a whole, both terrestrial and aquatic, in its function and structure. Contaminants at large contaminated sitesoften share critical properties such as high acute and/or chronic toxicity, high environmental persistence, often high mobilityleading to contamination of groundwater, and high lipophilicity leading to bioaccumulation in food webs. Contaminants presentat polluted sites occur as mixtures, therefore interactions between individual compounds are of importance.

The bioavailability is a key factor for ecotoxicological effects of contaminants. This is demonstrated by a case study onorganotins. Organotins belong to the most toxic pollutants known so far for aquatic life. Widespread contamination of harborsediments occurs globally due to the ongoing use of organotins in antifouling paints in large ships. In lake sediments, tributyl- andtriphenyltin are very persistent and bioavailable to biota even after a long time. Bioavailability of these compounds is dependenton pH and organic matter. Organotins are accumulated in sediments, but remobilization occurs when contaminated sedimentsa

nts to thee valuationa nough. As ac and landfilll xicity, andi es, whichc ntaminationb and presentg known.©

K ent; PAH

0

re disturbed and dredged.A key question in dealing with contaminated sites is the assessment and evaluation of the toxicity of contamina

nvironment. Usually, established OECD tests and whole effluent toxicity tests are performed for an ecotoxicological end for hazard assessment. However, these assays are often expensive, laborious and sometimes not sensitive eonsequence, we have used fast and reliable in vitro systems such as fish cell lines for the evaluation of sedimentseachates contaminated by polychlorinated hydrocarbons (PAH). Determination of cytotoxicity as a measure for acute tonduction of cytochrome P4501A (CYP1A) as a biomarker of exposure and toxicity were found to be important measuran be used for hazard and risk assessment. We have developed a concept for the ecotoxicological evaluation of PAH coased on induction equivalents, which can be applied for aquatic and terrestrial ecosystems. One of the key questionaps, however, includes the long-term chronic ecotoxicological effects on soil and aquatic biota, which are largely un2004 Elsevier Ireland Ltd. All rights reserved.

eywords:Ecotoxicology; Bioavailability; Organotins; Cytochrome P450 induction; Contaminated sites; Environmental risk assessm

E-mail addresses:[email protected], [email protected] (K. Fent).

300-483X/$ – see front matter © 2004 Elsevier Ireland Ltd. All rights reserved.doi:10.1016/j.tox.2004.06.060

224 K. Fent / Toxicology 205 (2004) 223–240

1. Ecotoxicology

In the last two decades ecotoxicology evolvedmainly from three different disciplines: toxicology,applied ecology and environmental chemistry. Eco-toxicology as an interdisciplinary environmental sci-ence deals with the interactions between environmen-tal chemicals and biota, thereby focusing on adverseeffects at different levels of biological organisation.Toxic effects of anthropogenic compounds in biotaand ecosystems are investigated in close connectionto their environmental chemistry and fate in the en-vironment. The bioavailability of chemicals, which isdependent on biogeochemical processes, is an impor-tant factor often neglected in ecotoxicological evalua-tion and hazard assessment. The bioavailable fractionis the critical parameter for uptake and ultimately forthe concentration at the target sites in organisms, whichis the critical parameter for toxicity (Fig. 1). Ecotox-icological research on selected pollutants requires aninterdisciplinary effort, considering physicochemical,molecular, toxicological, physiological and ecologicalprocesses. Whereas practical aspects of ecotoxicologyare mainly focused on regulatory issues (registrationof chemicals), and thus to testing of chemicals in stan-dardized tests, the focus of ecotoxicological research isaimed at an understanding of toxicological phenomenain a variety of biota, populations and the ecosystem as a

F lable fr moleculare anism

whole. Thereby, diverse aspects such as mechanisms oftoxic action and ecological processes in contaminatedsystems are regarded (Fent, 2003).

Ecotoxicological studies may also focus on ecolog-ical and toxicological effects observed in the field inretrospective studies, whereby a causative correlationbetween effects and chemical residue analysis is, how-ever, often difficult to establish. Ecological investiga-tions such as biomonitoring studies alone do not havesufficient resolving power to identify causative agents.Likewise, chemical analysis of pollutants in ecosys-tems alone cannot provide evidence for toxicologicalconsequences in biota. Only an integrated approachconsidering environmental chemical, toxicological andecological concepts may be suitable for understandingecotoxicological effects in contaminated ecosystems(Fent, 2003). One strategy to at least assess the contam-ination and its potential effects is the use of biomarkersin ecological surveys to verify the bioavailability andpresence of relevant concentrations in biota (Bucheliand Fent, 1996). A selection of the type of biomarkerallows a gross discrimination between certain groupsof contaminants, e.g. polyaromatic and dioxin-like pol-lutants versus heavy metals, as well as toxicologicalmechanisms or biological functions affected, e.g. geno-toxicity versus neurotoxicity or reproductive effects.

A more prospective approach is based on investi-gations of potential toxicological effects in laboratory

ig. 1. Ecotoxicological effects are dependent on the bioavaiffects that propagate to a variety of toxic manifestations in org

action of pollutants. Concentrations at the target sites induces.

K. Fent / Toxicology 205 (2004) 223–240 225

assays that may be used for extrapolation to the field.Bioassays play an important role in this process, how-ever, more comprehensive studies on contaminatedsystems and ecotoxicological processes are needed inaddition. Often, bioassays do not consider the processesin ecosystems, and neglect environmental factors thatinfluence toxicity. However, they are valuable tools inthe characterisation of the toxic action of chemicals,and in the understanding the associated toxicity. De-spite the usefulness of these tools, it should be notedthat the multitude of chemicals in ecosystems, speciesdiversity, biological and ecological functions and struc-tures makes extrapolations necessary for estimatingpossible effects in ecosystems. An important task there-fore is the improvement of the predictive power andquality of experimental systems and risk assessmentmodels.

In ecotoxicological research cellular effect studiesincluding knowledge of mechanisms of toxic action areas important as studies in laboratory species, becausethe primary interaction between chemicals and biotaoccurs at the surface of or in cells (Fig. 1). Whetherchemical-induced alterations in cell structure and phys-iology will develop into an adverse toxic effect dependson many parameters, including adaptive responses. Acellular effect is often, but not necessarily, deterministicfor adverse effects at higher levels of biological organi-zation. Factors such as compensatory mechanisms andthe presence of indirect effects may influence the rele-vance of the cellular toxicological response for overalle ela-t xic-i co-t mayu antf alth,a ogyp otox-i dat-i ogi-c theo ed,w icale msf tantsh asg isp

1.1. Ecotoxicological evaluation of contaminatedsites

A frequently used approach to evaluate ecologicalimpacts of contaminated sites is field monitoring. Thisincludes structural and functional measurements of res-ident biota, and sometimes in situ (caged) studies. Anecotoxicological evaluation usually consists of a mul-tispecies testing concept or often tiered approach, inwhich both environmental chemical properties as wellas effluents, extracts, or fractionated extracts are beingassessed. In contaminated aquatic sites, structural mea-sures include algal and aquatic plant density, biomass,benthic macroinvertebrate and fish surveys. In the ter-restrial part of contaminated sites, structural parameterssuch as plant density, soil invertebrate and biomass aredetermined. Structural indices include total abundance,taxa richness, diversity indices and various ratios of dif-ferent taxonomic groupings. Functional measurementsmonitor rate processes over time such as algal car-bon uptake, photosynthesis, and rates of reproductionor growth. Monitoring should be complemented within situ or experimental whole effluent toxicity assess-ments aimed at gaining knowledge about the ecotoxi-cological potential of pollutants at contaminated sitesand about cause relationships. In situ toxicity testing asan experimental approach can be used to measure vari-ation in exposure under actual receiving system condi-tions. In this approach, continuous discharge exposurescan be addressed, unlike grab samples, in which the ef-f plesa sess-m say-d bea ites(

1

glec t stan-d thee is am sedb er toh referst ts att prob-

cotoxicological effects of a given chemical. The rion between cellular toxicological responses to toty at higher biological levels is a key question in eoxicology. The hypothesis that cellular changesltimately influence biological parameters import

or populations such as growth, development, hend reproduction is obvious. Hence, cellular toxicolrovides an essential concept in understanding ec

cological processes, as it plays a key role in elucing toxic modes of action, and diagnoses toxicolal effects at higher biological levels, but it is notnly sufficient one. Its value will be strongly increashen it can be integrated more closely with ecologffects. Here, the applicability of in vitro cell syste

or the assessment of contaminated sites with polluaving different toxicological modes of action sucheneral (narcotic) toxicity and dioxin-like toxicityresented.

ects at one point in time or in 24 h composite samre assessed. In addition, whole effluent tests (asent of water samples or their extracts), and bioasirected identification of toxicants have proven ton useful tool in the evaluation of contaminated sBrack et al., 1999; Traunspurger et al., 1997).

.2. Ecological hazard and risk assessment

Whereas the ecotoxicological activity of sinhemicals can reasonably be assessed by currenardized ecotoxicological tests, the evaluation ofcotoxicological potential of contaminated sitesuch more difficult task. Environmental hazard poy a contaminated site is a source of potential dangumans and the environment. Hazard assessment

o evaluation of inherent properties of contaminanhese sites to cause harm. Risk is defined as the


Fig. 2. Environmental risk assessment. PEC, predicted environmen-tal concentration; PNEC, predicted no observed effect concentra-tion. Risk analysis is necessary when PEC/PNEC >1. Adapted fromECETOC (1993).

ability that a hazard will be realized. In a first phaseof an ecological risk assessment, in which sources andcontaminants of potential concern are identified, toxi-city testing is applied. In this regard, relatively simpleecotoxicity tests serve to identify hazard, and thus fita first stage of an eco(toxico)logical risk assessment,the hazard identification. The aim is to search for po-tential causal relationships among contaminants, re-ceptors and ecotoxicological endpoints. Further stagesof risk assessment—exposure and effects assessmentthen risk characterization—require additional informa-tion and can include other tools such as measures ofthe health of residents at contaminated sites, exposedpopulations of animals and plants, longer-term labo-ratory or field bioassays, toxicity identification evalu-ations (TIE) approaches, etc. Thereby, analysis of therelationship between contaminants and both laboratoryand field effects data is crucial. Finally, risk character-ization builds upon the results of the analysis phase todevelop an estimate of risk.Fig. 2illustrates a concept,often applied in chemicals risk assessment, which canalso be used for contaminated sites.

Often, leachates from sediments, soil and ground-water at contaminated sites are being tested in wholeeffluent toxicity tests. Whole effluent toxicity and sim-ilar toxicity tests integrate interactions among complex

mixtures of contaminants. They measure the total toxiceffect, regardless of physical and chemical composi-tion. This is a powerful integrative measure of the tox-icity of chemicals not being achievable by analyticalchemical measurements. Usually, direct toxic effectson survival, growth or reproduction are being assessedusing bacteria, algae or periphyton, water flea and fish.Although these tests have limitations and disadvan-tages such as not necessarily being environmentallyrealistic, they give important hints to the ecotoxico-logical potential of contaminated environmental mediaand can be adjusted based on site- and situation-specificconditions to be more predictive.

Often, however, there is a discrepancy between re-sults derived from both standardized laboratory testsand whole effluent tests compared to known biologi-cal impacts. For example, a set of 250 discharges fromcontaminated river systems across the United Stateswere tested in standardized Daphnia and fathead min-nows tests, whole effluent toxicity tests, and instreambiological condition as measured by benthic macroin-vertebrate assessments (species composition and abun-dance). The results indicated that whole effluent tox-icity testing was more predictive of biological effectsin the rivers, if several tests addressing different typesof endpoints were used (Diamond and Daley, 2000).Fish acute and chronic endpoints were most related toinstream condition, but no one endpoint was capableof accurately reflecting conditions of all discharges.This indicates that whole effluent toxicity testing is at ono ita-t lity,s toryt ef-fl dK oodp torsog tyr ares tivep n av di-v areu

s ther spite

,ool, which is useful for ecotoxicological evaluatif contaminated sites, but as all tools, has its lim

ions and imperfection. They include test variabipecies differences and extrapolations from laborao the field. For instance, it was found that wholeuent tests underestimated field effects (Clements aniffney, 1994), in other cases these tests were gredictors of fish response, but were poor predicf invertebrate response (Birge et al., 1989). With re-ard to reliability in predicting biological communiesponses laboratory single-species toxicity testsuggested, in a majority of cases, reliable qualitaredictors. The significance will be increased wheariety of different assays using different biota oferse evolutionary levels and ecological functionsed.

Laboratory species are generally not the same aesident species in the field aimed at protecting. De


this fact, risk assessment is based on laboratory testsfor mainly practical reasons (Burton, 1992; Linthurstset al., 1995). Extrapolating effects of toxicants froma limited number of test species to ecosystems as awhole is a difficult, but essential part of environmen-tal risk assessment (Koller et al., 2000). If toxicity testresults are available for very few species, the lowesttoxicity value is divided by an application factor orsafety factor that varies from 10 to 1000, dependingon the number of species tested and whether the end-point is based on acute mortality or effects (LC50 orEC50), or chronic no-observed-effect concentrations(NOEC). A comparison of 248 studies on 34 substancesinvolving both model ecosystems studies and chronicsingle-species studies indicate that an assessment fac-tor of 8 would be appropriate in the extrapolationfrom the lowest chronic singe-species NOEC-value ina model ecosystem (ECETOC, 1993). Applying safetyfactors in the range of 10–1000 for extrapolation fromthe laboratory to field is therefore appropriate. Thisrisk assessment concept using the ratio between pre-dicted exposure concentration (PEC) and predicted noobserved effect concentration (PNEC), as outlined inFig. 2, is widely accepted (EC, 1994, 1996; ECETOC,1993). However, it has also its limitations and resultsof such analyses should be judged and interpreted withnecessary caution with respect to ecological relevance(Koller et al., 2000).

Analyses using ecotoxicological tests with labora-tory species may provide an uncertain level of pro-t oticf ofc varyi sts.S ner-a entt io-c s isn onp dif-f teds o oc-c sucha or-g elyp ch-a ation(

In the following, two case studies are presented, inwhich the applicability of fast and reliable in vitro cellculture systems for the assessment of chemicals havingdifferent modes of actions, such as general toxicity anddioxin-like toxicity, is demonstrated. Such chemicalsare often present at contaminated sites. In addition, theimportance of the bioavailability of contaminants fortheir ecotoxicological effects is underlined.

2. Case study: organotins

2.1. Contamination of harbor sediments

Organotin compounds are among the most haz-ardous pollutants known so far in aquatic ecosystems(Fent, 1996). Tributyltin (TBT) is of particular impor-tance because of its widespread use as biocide, namelyin antifouling paints on ships and in wood protection.Since the late 1970’s considerable quantities of TBTwere introduced into the aquatic environment and as aresult, widespread pollution of marine and freshwaterharbors and adjacent areas resulted. Organotin pollu-tion in the aquatic environment is of global concern.Due to the extreme toxicity and the ecotoxicologicalhazards associated with TBT in antifouling paints (Fentand Meier, 1992; Horiguchi et al., 1997) restrictions onits use have been implemented in many countries in themid to end 1980’s. As a consequence, TBT concentra-tions in harbor waters decreased significantly at manyl ;F sei l.,1i l.,1 seo se ina romp rva-t nts( urals noxics si se-q sh-w up tos esentl s for

ection for several reasons. First, abiotic and biactors influence the bioavailability and toxicityontaminants. Often, contaminant concentrations

n time, which is not regarded in laboratory teecond, mixtures of a set of contaminants as gelly found in contaminated sites may show differ

oxicity compared to single compounds. Third, boncentration and bioaccumulation in food webot regarded. Fourth, they do not reflect effectsopulations or community responses and may

er from the situation in the field. In contaminaites, adaptations of biota to contaminants can alsur. In general, adaptation has ecological costss energy consumption, which may reduce theanism’s fitness. Adaptive reactions include activumping out incoming toxicants, detoxification menisms, damage repair and avoidance of contaminHansen et al., 1999).

ocations in industrialized countries (Chau et al., 1997ent and Hunn, 1995), but this is not or less the ca

n other locations (Biselli et al., 2000; Takahashi et a999), in developing countries (Kannan et al., 1995) or

n marine mammals (Iwata et al., 1995; Kannan et a997; Tanabe, 1999). In spite of regulations, the releaf TBT into aquatic ecosystems persists due to its untifouling paints on large vessels, paint removal fleasure boats, application of TBT in wood prese

ion, and remobilization from contaminated sedimeFig. 3). Because TBT associates strongly with natorbents and because of its high persistence in aediments (Fent and Hunn, 1991), TBT accumulaten sediments for a long period of time. As a conuence, high levels of TBT have been found in freater and coastal sediments, with concentrationseveral mg/kg. Such contaminated sediments reprarge contaminated sites, but also important source


Fig. 3. Contamination of pleasure boat harbors in Germany in the North Sea and in the Baltic Sea. Harbor locations: Heiligh., Heilighafen;Warnemu 1, Warnemunde site 1; WM 2, Warnemunde site 2. Data afterBiselli et al. (2000).

long-term pollution of the aquatic environment dueto remobilization. As shown inFig. 3, high con-tamination of marine harbor waters and sedimentswith tributyltin occurred long after regulation of TBT-antifouling paints. In addition, alternative pesticidessuch as the s-triazine herbicide Irgarol 1051, which isused as a replacement for organotins, has been foundin considerable concentrations (Biselli et al., 2000).

Triphenyltin (TPT) has been used as co-toxicantwith TBT in some antifouling paints (Fent and Hunn,1991), although its major application lies in agricul-ture, where it is used as a fungicide for various cropsand enters aquatic ecosystems via leaching and runofffrom agricultural fields (Stab et al., 1996). In wastewa-ter and sewage sludge (Fent and Muller, 1991), as wellas in water and sediment, degradation products of TBT,dibutyltin (DBT) and monobutyltin (MBT), as wellas from TPT, diphenyltin (DPT) and monophenyltin(MPT), respectively, occur in addition to the parentcompound (Fent, 1996; Fent and Hunn, 1991).

Prior to implementation of the sales ban of TBT-containing antifouling paints in Switzerland in 1990,we have measured high concentrations of organotinsin sediments of freshwater pleasure boat harbors (Fentand Muller, 1991). Five years later, we have sampledthe harbors again and analysed the anoxic sedimentsfor organotins. Sediment core samples were taken in

1994 by a gravity corer from the pleasure boat harborLucerne in Lake Lucerne. Sediments were extrudedfrom Plexiglas liners in 1 cm slices and analyzed usinghigh resolution capillary gas chromatography accord-ing to the procedure described (Fent and Hunn, 1991).

Butyltin and phenyltin profiles in the sediment coreare depicted inFig. 4. Relatively high concentrations ofTBT concentrations were found in the upper 3–4 cm.The decrease below 4 cm mainly represents the loweruse of antifouling paints and associated accumulationof organotins. This and our former analysis with dat-ing of the sediment cores demonstrates that the sed-iment profiles reflect the historical use of organotincontaining antifouling paints and input into the sed-iment. These antifouling paints were introduced inthe early 1970s in Switzerland and their consumptionreached a peak in mid 1980s. Degradation productsDBT and MBT occur as well. The higher concentra-tions at 0.5–1.5 cm depth indicate increased degrada-tion, but in lower parts, the proportion of degrada-tion products remains low, indicating low degradationof TBT. Phenyltins were also present at a depth of1–4.5 cm at relatively low concentrations up to max-imal 6�g/kg (dry weight), which indicates recent useof TPT-containing antifouling paints for a short periodof time. These data clearly show that both TBT andTPT remain in the sediments of harbors for long time


Fig. 4. Butyltin (A) and phenyltin (B) concentrations (dry weight)in a habor sediment core taken in the pleasure boat harbor Lucerne,Switzerland, 4 years after implementation of TBT-antifouling paintsales ban. TBT, tributyltin; DBT, dibutyltin; MBT, monobutyltin;TPT, triphenyltin; DPT, diphenyltin; MPT, monophenyltin.

periods. Four years after the antifouling paint sales ban,however, concentrations start to decrease. As degrada-tion products do not increase significantly with increas-ing depth, degradation is very slow.

Sorption of TBT and TPT to sediments is fast, butreversible (Unger et al., 1988). The organic material innatural sorbents plays the pivotal role in the sorptionof hydrophobic TBT and TPT. Complexation of TBTand TPT cation with carboxylate and phenolate ligandsin organic material plays the dominant role. Besideshydrophobic interactions, adsorption of organotins toclay minerals was found at low pH by a cationic ex-change mechanisms (Hermosin et al., 1993). Organotincompounds will readily desorb at storm events andwhen dredging the sediments. Therefore any resus-pension of organotin contaminated sediments will leadto enhanced organotin concentrations in the overly-ing water making these compounds bioavailable toaquatic organisms. In addition, organotins in pore wa-ter are bioavailable to benthic organisms. This leadsto the conclusion that organotin-contaminated sedi-ments, representing large contamination sites, act asan organotin source for organisms living at or insediments.

2.2. Bioavailability

Ecotoxicological effects are largely dependent onthe bioavailability of the contaminants. The bioavail-ability of environmental chemicals can be studiede Sev-e icals minet anicc owb ionc axi-m1 ciesi eda manyc undb annotb ptakee ionc

omn and

xperimentally by assessing the bioconcentration.ral general factors such as hydrophobicity, chempeciation and organism-specific properties deterhe degree of bioconcentration. For neutral orgompounds of medium hydrophobicity and liotransformation potential, octanol–water partitoefficients (Kow) can serve as an estimate of the mum possible bioconcentration in organisms (Mackay,982). For chemicals that can form charged spe

n water, different bioconcentration of the chargnd uncharged species should be expected. Inases, however, simple partitioning of the compoetween the aqueous phase and the organisms ce assumed. Species-specific factors such as ufficiency, growth dilution, metabolism and excretan modify the degree of bioconcentration.

Organotin compounds differ in their properties freutral organic compounds. The speciation of TBT


TPT shows a strong pH-dependence (Arnold, 1998).Both compounds are present as cations at low pH andas hydroxides at higher pH. While they can also formcomplexes with other anions, these species are onlyof minor importance in typical freshwater systems.Therefore, only the hydroxide species, TBTOH andTPTOH, and the cations, TBT+ and TPT+, have tobe considered. Hydroxides and cations, however, ex-hibit very different partitioning and sorption behavior.The octanol–water distribution ratios (Dow) of TBT andTPT are more than an order of magnitude higher at pH 8than at pH 3. At pH 8, a logKow of TBTOH and TPTOHof 4.10 and 3.53, respectively, was determined (Arnold,1998). This is related to the fact that TBTOH and TP-TOH, but not TBT+ and TPT+, readily partition intothe octanol phase.

What is the influence of pH and humic acids on thebioconcentration of TBT and TPT in different aquaticbiota and at different conditions? InDaphnia magna,TBT and TPT bioconcentration was dependent on thepH dependent speciation (Fig. 5). Significantly higherbioconcentration occurred at pH 8 than at pH 6 or5, which is related to the fact that these compoundsmainly, but not exclusively, accumulate as hydroxideswhich are predominant at pH 8 (Fent and Looser, 1995;Looser et al., 1998). At pH 8, TBT occurs predom-inantly as TBTOH (95%), whereas the fraction ofTPTOH is more than 99%. At pH 5, TBT is presentprimarily as positively charged TBT+ (95%), whereasin the case of TPT, both species are present similarfF andT ,

F ent pH yltins 2–10. AF

using tetrabutyltin (TeBT) as reference compound thatmay only undergo hydrophobic partitioning. At bothpH, TPT bioconcentration was significantly higher ascompared to TBT. The major reason for this findingis the significant metabolism of TBT, but not TPT inChironomus riparius(Looser et al., 2000). In additionto the neutral TPTOH, the TPT+ is also taken up.

Dissolved organic matter such as humic substanceslead to a significant reduction in the bioavailabilitydue to hydrophobic sorption of organotins (Fent andLooser, 1995; Looser et al., 1998, 2000). The effect ofAldrich humic acids (AHA) on the bioconcentrationof TBT and TPT inDaphniaand fish yolk sac larvaeThymallusis presented inFig. 7. Aldrich humic acidsreduced the bioconcentration of TBT and TPT in thebenthic midge larvaeChironomus. These experimentsclearly demonstrate that the bioavailability of organ-otins is a function of pH and concentration of dissolvedorganic matter. In addition, they demonstrate that thebioavailability is a key parameter governing toxicity,as it goes along with the magnitude of bioaccumlationand hence, concentration at target sites.

2.3. Ecotoxicity

Organotins are extremely toxic to aquatic biota asdemonstrated for a variety of different organisms invivo and in vitro (Fent, 1996). Many ecotoxicologicalstudies on organisms of different evolutionary levelhave been reported (Alzieu, 2000; Alzieu and Heral,1 92;F 93;H 98

ractions (61 and 39%, respectively) (Arnold, 1998).ig. 6compares the bioconcentration of TBT, TPTeBT in Chironomus ripariuslarvae at pH 8 and 5

ig. 5. Bioaccumulation of tributyltin inDaphnia magnaat differpecies at pH 2–10; (C) octanol–water distribution rates at pH

: (A) bioconcentration at pH 6 and 8; (B) occurrence of tributfterent and Looser (1995)andFent (1995).

984; Bryan et al., 1989; Fent and Meier, 19ioramonti et al., 1997; Hamasaki et al., 19origuchi et al., 1997; Mathiessen and Gibbs, 19).


Fig. 6. Bioaccumulation of tributyltin (A); triphenyltin (B); andtetrabutytlin (C) inChironomus ripariuslarvae at pH 8 and 5. Mean± S.D. of 50–60 larvae of five to six experiments:cb, concentrationin Chironumus; cw, concentration in exposure water. AfterLooseret al. (1998).

Lowest toxic concentrations are in the range of1–10 ng/L. Marine gastropods and oysters are amongthe most susceptible organisms, but fish are affectedas well, although at concentrations of 1–10�g/L.However, the long-term ecotoxicological effects oforganotins on the structure and function of aquaticecosystems are still not well understood, particularlywith respect to biomagnification in food webs (Gurugeet al., 1996; Jak et al., 1998; Kannan et al., 1996; Kimet al., 1996; Stab et al., 1996).

TBT and TPT act via different modes of action.Perturbation of calcium homeostasis, inhibition of mi-tochondrial oxidative phosphorylation and ATP syn-thesis, inhibition of photophosphorylation in chloro-plasts, genotoxicity and DNA damage, and inhibitionof enzymes such as ATPases and cytochrome P450monooxygenases (CYP) are among the key biochemi-cal processes (Fent, 1998). Therefore, key biochemicalprocesses are affected. Inhibition of CYP enzymes, inparticular the aromatase, responsible for the conversionof testosterone to estradiol, seems to be responsiblefor the masculinization of female marine gastropods,of which over 100 species are known to be affectedby TBT. The females develop male reproduction or-gans at trace concentrations of a few ng/L TBT, mostprobably due to inhibition of aromatase and associ-ated disturbance of steroid metabolism (Bettin et al.,1996; Oehlmann and Bettin, 1994). Inhibition of CYPby TBT and TPT has also been demonstrated in fishin vivo (Fent and Stegeman, 1993) and in vitro (Fenta -h ed inmt thatt estisw ed at1

eco-t or-g per-f rna-t g ofc sed( rep shc icc cel-

nd Stegeman, 1991; Fent et al., 1998). Recently, inibition of aromatase has also been demonstratollusks (Morcillo and Porte, 1997), and mammals. A

wo-generation reproductive study of TBT showedhe male reproductive system of rats is affected. Teights, spermatid and sperm count were reduc25 mg TBT/kg (Omura et al., 2001).

One of the most commonly used bioassay inoxicology for assessing acute toxicity to aquaticanisms is the acute lethality test in fish, usually

ormed according to OECD guidelines. As an alteive or a supplementary bioassay for toxicity rankinhemicals, in vitro cytotoxicity assays have been uBabich and Borenfreund, 1988, 1991; Bruschweilet al., 1995; Fent and Hunn, 1996). Monitoring ofollution can be determined in vitro by using fiell lines (Fent, 2001). Interaction of anthropogenhemicals with biota takes place first at the


Fig. 7. Bioaccumulation of tributyltin and triphenyltin inChironomus ripariuslarvae at pH 8 and 5 in the absence or presence of 23 mg C/LAldrich humic acids. Bioconcentration factor ascb/cw. cb, concentration inChironumus; cw, concentration in exposure water. Percentage oforganotin species at respective pH is also given above columns. Data afterLooser et al. (2000).

lular level making cellular responses not only thefirst manifestation of toxicity, but also suitable toolsfor the early and sensitive detection of chemicalexposures.

Cytotoxicity was found in several studies to be cor-related with in vivo acute fish toxicity (Castano et al.,1995; Segner, 1998) in different cell lines, which wasalso found for organotins. Moreover, cytotoxicity oforganotins was correlated to the octanol–water parti-tion coefficient (Bruschweiler et al., 1995; Fent, 1996).These studies lead to the conclusion that fish cell line-based cytotoxicity assays are important in contributingto the evaluation of the environmental hazard of chem-icals and environmental samples. Ideally, they shouldbe used in combination with other ecotoxicological testsystems covering organisms of different evolutionarylevels and ecology.

3. Case study: polycyclic aromatichydrocarbons and related cytochrome P450inducers

Environmental compounds that elicit toxic effectssimilar to that of polychlorinated dioxins (dioxin-likechemicals) are of concern due to their high toxicity

to organisms of different evolutionary levels. Severaltoxic and biochemical effects of such chemicals are me-diated through the aryl hydrocarbon receptor (AHR),which is highly conserved in evolution and found in nu-merous taxa (Hahn, 1998). AHR is a ligand-dependenttranscription factor located in the cytosol of the cell(Hahn et al., 1993). The ligands for AHR are hydropho-bic aromatic compounds with planar structure of par-ticular size, which fits the binding site. Upon bindingof an agonist AHR is translocated to the nucleus, whereit forms a heterodimer with the AHR nuclear translo-cator protein and binds to dioxin-responsive elementsin the promoter region of certain genes, which are thenupregulated. One of the expressed genes is cytochromeP4501A (CYP1A), a key enzyme in the metabolism ofxenobiotics. It has been shown that the binding strengthof a ligand to the AHR is roughly directly proportionalto the enhanced gene transcription and associated tox-icity.

3.1. An induction equivalency concept for thetoxicity assessment of polycyclic aromatichydrocarbons

CYP1A induction is a sensitive and specific adap-tive response of organisms exposed to an important


series of environmental pollutants such as planar con-geners of polychlorinated dibenzodioxins and diben-zofurans (PCDD, PCDF), polychlorinated biphenyls(PCB), several polycyclic aromatic hydrocarbons(PAH), polychlorinated naphtalenes and related com-pounds (Bucheli and Fent, 1996; Stegeman and Hahn,1994). CYP1A induction is used as a biomarker indica-tive for exposure to such pollutants in organisms suchas fish and fish cell systems. Fish cells contain an AHR,which is very similar to mammalian AHR, but less ex-pressed (Hahn, 1998). We were studying the inductionpotential of a diverse category of environmental chem-icals and environmental samples in the fish cell linePLHC-1 to determine their induction potential and tox-icity. Permanent cell lines tend to lose their metabolicactivity for detoxification of xenobiotics. However,several fish cell lines from the liver such as PLHC-1 andRTL-W1, and some fibroblast-like cells like RTG-2,BF-2, FHM retain at least a basic cytochrome P450 de-pendent monooxygenase activity (Fent, 2001). The per-manent cell line PLHC-1 derived from the topminnow,Poeciliopsis lucida, after exposure to the carcinogen7,12-dimethylbenzanthracene (Hightower and Renfro,1988) has been applied in various types of studies.This and other fish cell lines has been demonstratedto bear metabolic activity (Babich and Borenfreund,1991) and to contain an aryl hydrocarbon receptor(AHR) (Hahn, 1998). Recently, we demonstrated thatalso very low levels of estrogen receptor mRNA areexpressed (Haugg et al., 2000). This cell line hast omeP talp Bs)( ahne( es( n-m l.,1 esi eda ark-e 5;S r-m yre-s on-t ternb l.,1

Cost-effective bioassays for aquatic systems that in-tegrate biological effects of complex mixtures of PAHand other pollutants are needed. An estimation of thetoxic potential of environmental mixtures has beenused for halogenated aromatic hydrocarbons (HAH) byapplying the toxic equivalency (TEQ) concept (Safe,1994; Zabel et al., 1996). Thereby, the toxic or bio-chemical potency of a compound or mixture of thesepollutants in biological systems is compared with thepotency of the most toxic compound, e.g. 2,3,7,8-TCDD. TEQs are the sum of the products of the con-centrations of the pollutants and their toxic equivalencyfactors (TEF), which are estimated relative to TCDD.This concept is based on in vivo and in vitro studies withcultured mammalian cells that indicate that these com-pounds cause similar effects, but differ in potency. TheTEF concept has also been developed by other AHR-mediated endpoints such as receptor binding, and fishearly life stage mortality. By using this concept, envi-ronmental concentrations can be expressed in terms ofbiological responses that are bioassay derived.

Polycyclic aromatic hydrocarbons (PAHs) haveonly rarely been analysed for CYP1A inductionin fish in vitro systems. We used PLHC-1 cellsto derive fish specific TEFs, or as better called,induction equivalency factors (IEFs), for 19 PAHs(Fent and Batscher, 2000), 12 nitrated PAHs (NPAH)and 12 azaarenes (Fig. 8) (Jung et al., 2001). Therelative CYP1A induction potencies, determined asethoxyresorufinO-deethylase (EROD) activity, andt sixb atedlB edw thene,b onw Allo ledt ithr ltingiT onc nd( ulisa a-l ont s,t z[

he capacity to induce the expression of cytochr4501A (CYP1A) after exposure to environmenollutants such as polychlorinated biphenyls (PCBruschweier et al., 1996; Clemons et al., 1996; Ht al., 1993, 1996; Hestermann et al., 2000), PAHsFent and Batscher, 2000) nitrated PAH and azaarenJung et al., 2001) as well as to extracts of enviroental matrices (Villeneuve et al., 1997; Willet et a997). As an indicator of toxicant-induced chang

n biological systems, CYP1A induction is regards one of the most sensitive and specific biomrs for such contaminants (Bucheli and Fent, 199tegeman and Hahn, 1994). CYP1A is either deteined by its enzyme activity such as the ethox

orufinO-deethylase (EROD), or by the protein cent using immunocytochemical methods (Weslots, ELISA) (Bruschweier et al., 1996; Hahn et a993).

he cytotoxicities of 19 compounds with one toenzene rings, mixtures of PAHs, and contamin

andfill leachates have been determined (Fent andatscher, 2000). No CYP1A induction was observith benzene, naphthalene, anthracene, acenaphenzo[g, h, i]perylene and fluorene, but low inductias found with fluoranthene and phenanthrene.ther PAHs with three and more benzene rings

o a concentration-related induction of CYP1A, webound decreases at high concentrations resun bell-shaped concentration-activity curves (Fig. 8).he inhibition at higher concentrations is basedompetitive inhibition by the inducing compouFent and Batscher, 2000; Hahn et al., 1996; Petrnd Brunce, 1999). Fish-related induction equiv

ency factors (IEFs) were estimated for all PAHshe basis of EC50 values of their EROD activitiehereby taking the most active compound, dibena,


Fig. 8. Induction of CYP1A in PLHC-1 cells determined by catalytic activity. Concentration–EROD activity curves of representative polycyclicaromatic hydrocarbons in PLHC-1 cells. DaiP, dibenz[a, i]pyrene; Ch, chrysene; BaP, benzo[a]pyrene; BaA, benz[a]anthracene; Na, naphthalene.Adapted fromFent and Batscher (2000).

h]anthracene, as the reference compound for the IEFs(Table 1). The following order of decreasing IEFs wasfound: dibenz[a, h]anthracene > dibenzo[a, i]pyrene> benzo[k]fluoranthene > 3-methylcholanthrene> benzo[a]pyrene > benzo[e]pyrene > chrysene

Table 1Induction equivalency factors (IEF) of PAH in PLHC-1 cells

Compound IEF

Dibenz[a, h]anthracene 1.0Dibenzo[a, i]pyrene 0.42Benzo[k]fluoranthene 0.303-Methylcholanthrene 0.13Benzo[a]pyrene 0.050Benzo[e]pyrenea 0.040Chrysene 0.0297,12-Dimethylbenz[a]anthracene 0.011Perylenea 0.0091Benz[a]anthracene 0.0053Pyrene 0.0022Benzo[g, h, i]perylene –Fluoranthene –Phenanthrene –Anthracene –Acenaphthene –Fluorene –Naphthalene –Benzene –

Data afterFent and Batscher (2000).a Substances with low maximal EROD activities; (–) no signifi-

cant EROD induction.

> 7,12-dimethylbenz[a]anthracene > perylene >benz[a]anthracene > pyrene. In contrast to the ERODactivity showing a bell-shaped concentration-activitycurve due to competitive substrate inhibition, theimmunodetectable protein content determined byELISA showed a concentration-dependent increase(Fent and Batscher, 2000; Hestermann et al., 2000).

3.2. Nitrated polycyclic aromatic hydrocarbonsand effects of compound mixtures

Nitrated polycyclic aromatic hydrocarbons (NPAH)andN-heterocyclic aromatic hydrocarbons (azaarenes)are as ubiquitous in the environment as their parentPAH compounds, although occurring at lower concen-trations. The toxicological importance of NPAH andazaarenes is based on their mutagenic and carcino-genic potential. Azaarenes possess a higher solubilityand mobility in the environment than PAH. However,very little is known about the toxicity and CYP1Ainduction of NPAH and azaarenes in fish. We deter-mined the cytotoxicities and relative CYP1A induc-tion potencies of 12 NPAH and 12 azaarenes, deter-mined as neutral red uptake and EROD activity, re-spectively (Jung et al., 2001). Additionally, CYP1A en-zyme protein was determined by ELISA for two NPAH,azaarenes, PAH and binary mixtures. Compared tothe structurally analogous PAH, 2-nitronaphthalene, 9-


nitroanthracene, 3-nitrofluoranthene, benzo[a]acridineand benzo[h]quinoline revealed higher induction po-tencies, whereas the other compounds showed similaror less activity. The induction potency was highly de-pendent on the compound’s structural properties, re-flected by significant correlations between the half-maximal EROD induction (−log EC50) and the molec-ular descriptors lipophilicity (logKow) and maximalmolecular length (Lmax) (Fig. 9). The CYP1A induc-tion potencies of NPAH and azaarenes suggest that theircontribution to the overall CYP1A induction potenciesin PAH-contaminated sites have to be taken into ac-count.

In contaminated sites, pollutants occur generallyas mixtures, but little is known about the interac-tion of compounds. We found that the interaction ofPAHs in mixtures of up to eight individual compoundswas additive based on their EROD activities (Fentand Batscher, 2000). This was also found for otherAHR-inducing compounds having the same mode oftoxic action. Binary mixtures of 6-nitrochrysene +benzo[a]anthracene, 6-nitrochrysene + benzo[a]acri-dine and benzo[a]acridine + benzo[a]anthraceneshowed an additive interaction (Jung et al., 2001).Based on these findings an evaluation of the CYP1Ainduction potential of environmental samples can beperformed taking into account an additive behavior ofindividual compounds. The advantage of such in vitrosystems is that the whole mixture of contaminants isevaluated, therefore the results of these assays can bei

3 gfi

l iso ac-c ma-r pol-l m,e ilarc ef-f e fo-c ando tilityo xi-c ands

Fig. 9. Structure–activity relationships of PAH, nitrated polycyclicaromatic hydrocarbons and azaarenes. Shown are correlations be-tween octanol–water partition coefficient, logKow (A), and maxi-mal molecular length of the molecule,Lmax, respectively (B), with−log EC50 derived from in vitro studies in PLHC-1 cells. Arrowsindicate minimal EC50 values used in the regression analyses forcompounds without full concentration response curves. Correlationcoefficientsr2 = 0.83 (logKow) and 0.73 (Lmax), respectively. AfterJung et al. (2001).

ncluded in the ecological risk assessment.

.3. Assessment of PAH-contaminated sites usinsh cell lines

Pollution of aquatic environments by mineral oif global concern. The frequently occurring tankeridents are associated with severe acute toxicity inine and freshwater ecosystems. However, chronicution by low concentrations of mineral oil, petroleutc. from a variety of different sources is of simoncern, but little is known about the long-termects in pelagic and benthic ecosystems. Here, wus on a severely polluted river system in Estoniather contaminated sites and demonstrate the uf fish cell line in vitro systems for rapid ecotoological assessment of PAH-contaminated waterediment.


Among watersheds at the oil shale area of Esto-nia, River Purtse with its tributaries, is heavily pollutedwith wastewater from coke-ash dumps of oil shale pro-cessing plants and with oil shale drainage water fromunderground mines and open-cast pits. Impacts havebeen recorded not only in this river and its tributaries,but also the Gulf of Finland (Huuskonen et al., 2000).One of the benefits of using fish cells for assessmentof mineral oil pollution is rapidity, which is necessaryin making predictions based on a larger data set in riskassessment. The evaluation of the CYP1A inductionpotential of sediments from River Purtse and River Ko-htla, Estonia, was performed in determining CYP1Ainduction and porphyrin accumulation of hexane ex-tracts (Huuskonen et al., 2000). In contrast to aqueousextracts with which the bioavailable fraction of pollu-tants is assessed, organic solvent extracts give a worstcase estimate. All sediment extracts led to CYP1A in-duction and porphyrin accumulation in cells. The mostactive sediments originated near the oil shale process-ing plants, which contained very high concentrations ofPAH. The biological potency in cells and PAH contami-nation of the samples showed the same rank order, withsome exceptions. This indicates that the in vitro cellsystem provides a sensitive bioanalytical tool for sedi-ment analysis contaminated with PAH-type pollutants.A recent study using a reporter gene assay demon-strated a similar usefulness in environmental risk as-sessment (Vondracek et al., 2001).

We were using the PLHC-1 cell line also for the as-s po-tB pt,t ex-t andm atesc se-d red,E Qs)c uredP ancyb thec een> Qs,wB in-d ncy,o , re-

Fig. 10. CYP1A induction potency of organic landfill leachate ex-tracts (S1–S6), contaminated with a variety of PAH. Leachates, butnot uncontaminated reference extracts (Ref), lead to a concentration-related induction of CYP1A activity in PLHC-1 cells. As with purecompounds, inhibition of EROD activity occurred at high concen-trations. AfterFent and Batscher (2000).

constituted mixtures, consisting of eight different PAHsin concentrations found in landfill leachates, wereanalysed. The good correlation between the bioassay-derived and calculated IEQs indicates that the PAHmixture in reconstituted and natural samples showedan additive behavior. The demonstration of additive in-teractions is important in hazard and risk assessmentof PAH-contaminated environmental samples and sup-ports the additive model for the prediction of IEQs(Fent and Batscher, 2000). It is impossible to iden-tify all of the compounds in these landfill leachatesby chemical analysis and therefore unidentified com-pounds will increase the bioassay-derived IEQs com-pared to calculated IEQs. Hence, the increased IEQsare likely caused by additional unknown inducing com-pounds including unidentified PAHs, substituted PAHs,PCBs, PCDDs, PCDFs or dibenzothiophenes probablypresent in the landfill leachates. Similar observationshave been made in other studies using mammalian celllines (Willet et al., 1997).

In conclusion, the additive behavior of PAHs isimportant for the successful application of the IEQconcept for hazard and risk assessment of PAH-contaminated sites. The demonstration of the induc-tion potential of landfill leachates demonstrates the

essment of the cytotoxicity and CYP1A inductionential PAH-contaminated landfill leachates (Fig. 10).y applying the induction equivalency (IEQ) conce

he ecotoxicological potential of several organicracts of landfill leachates, contaminated sedimentsotorway-runoffs were assessed. In landfill leach

ontaminated with a large variety of PAHs, doependent EROD induction curves were measuC50 values derived and induction equivalents (IEalculated based on both the analytically measAHs, and cell-derived IEFs. There was a discrepetween the determined IEQs in PLHC-1 cells andalculated IEQs: bioassay derived IEQs were betw4 and 112 times higher than the calculated IEhich are based on chemical analyses of PAH (Fent andatscher, 2000). In order to test whether additionalucing compounds are responsible for this discrepar whether the PAH mixture showed a synergism


advantage of bioanalytical tools such as the PLHC-1 cell system over chemical determinations alonefor the ecotoxicological evaluation of environmentalsamples.

4. Conclusions

There are many important classes of contaminantsat large contamination sites that are not studied in de-tail, partly due to the lack of suitable instrumentalanalytical techniques. As chemical analysis is aimedat identifying and quantifying specific environmentalchemicals, others not covered by the analytical tech-nique are neglected. Another problem is the interac-tion of chemicals in complex mixtures, which is notyet understood. Currently, environmental contamina-tion is assessed mainly by chemical analysis. How-ever, concentrations analysed provide only part of theknowledge necessary to evaluate and assess the toxicpotential of pollutants at contaminated sites for wildlifeand humans. This is partly because the bioavailabilityof the compounds is not considered and because eachof the compounds has different biological activities.Moreover, the complex interactions between differentenvironmental chemicals are not understood and con-sidered when hazard assessments and predictions ofpossible ecotoxicological effects are made based onconcentrations alone.

The use of selective bioanalytical tools, particularlyi entt oas-s e oft ane alsi t actt ratet nt int lativep tiont giv-i basedo % oft ECv thei osta fs equiv

alents, and can be integrated into a hazard and riskassessment procedure.

Our studies with samples from contaminated sitesdemonstrate that the fish cell in vitro system serves asan integrative bioanalytical tool in the ecotoxicologi-cal evaluation of samples contaminated with CYP1Ainducing compounds (Fent and Batscher, 2000;Huuskonen et al., 2000). They are valuable tools in thefirst evaluation of contaminated environmental sam-ples. Bioassays assess the general toxicity (cyotoxicity)and the activity of compounds that act through a spe-cific receptor-mediated mechanism of action (CYP1Ainduction). Generally, data from bioassays have eco-toxicological relevance, because they represent an in-tegrated biological response, however, there are alsolimitations. In vitro assays do not account for the bioki-netics, tissue distribution and biotransformation thatmay occur in vivo. Bioassays cannot identify the in-dividual compounds causing the response, but givean overall activity of the sample, which is of coursethe most relevant information for environmental riskassessment. Therefore, ecotoxicological evaluation ofcontaminated sites should be based on a set of differentin vitro and relatively simple in vivo bioassays besideschemical analysis.

Ecotoxicological effects of chemicals are dependenton the exposure and bioavailability of compounds, up-take and metabolism, intracellular concentration, modeof toxic action and balance between toxicity and protec-tive cellular responses. In ecotoxicology, not only thev o theb dif-f mon-s ntalf orp-t me-t uldb . It ish teds ly one toa

A

nd,a an-

n connection with chemical analysis, can circumvhese limitations. In vitro cell systems used as biays offer a rapid and sensitive solution to somhe limitations of chemical analysis. They enablestimate of the total biological activity of chemic

n environmental matrices or extracts thereof thahrough the same mode of action. They also integhe interaction among the different chemicals presehe tested sample from contaminated sites. The reotencies of samples are usually calculated in rela

o the amount of standard, or reference compoundng the same response as the sample, commonlyn the amount of sample needed to produce 50

he maximal standard response and calculated as50alue. In case of AHR-mediated activity, we usednduction equivalency concept normalized to the mctive compound, dibenz[a,h]anthracene. Activities oamples are then expressed as bioassay-derived
-
ast variety of organisms has to be regarded, but alsioavailablity of toxicants, which is dependent on

erent environmental processes. This has been detrated for organotin compounds, where environmeactors such as pH, humic acids and reversible sion to sediments were found to be important paraers. The aspect of bioavailability of toxicants shoe better integrated in risk assessment conceptsoped that in future the evaluation of contaminaites as well as hazard and risk assessment will recotoxicological meaningful bioassays in additionnalytical chemical measurements alone.

cknowledgement

I thank the 3R Research Foundation Switzerland the Umweltbundesamt Berlin, Germany, for fin


cial support, and my students R. Batscher, D. Jung andP. Looser for collaboration.

References

Alzieu, C., 2000. Impact of tributyltin on marine invertebrates. Eco-toxicology 9, 71–76.

Alzieu, C., Heral, M., 1984. Ecotoxicological effects of organotincompounds on oyster culture. Ecotoxicol. Test. Mar. Environ. 2,187–196.

Arnold, C.G., 1998. Triorganotin compounds in natural waters andsediments: aqueous speciation and sorption mechanisms. Ph.D.thesis, Swiss Federal Institute of Technology (ETH), Zurich.

Babich, H., Borenfreund, E., 1988. Structure–activity relationshipsfor diorganotins, chlorinated benzenes, and chlorinated anilinesestablished with bluegill sunfish BF-2 cells. Fundam. Appl. Tox-icol. 10, 295–301.

Babich, H., Borenfreund, E., 1991. Cytotoxicity and genotoxicity as-says with cultured fish cells: a review. Toxicol. In Vitro 5, 91–100.

Bettin, C., Oehlmann, J., Stroben, E., 1996. TBT-induced imposexin marine neogastropods is mediated by an increasing androgenlevel. Helgol. Meeresunters 50, 299–317.

Birge, M.T., Black, H.A., Short, T.M., 1989. A comparative ecolog-ical and toxicological investigation of a secondary waste-watertreatment plant effluent and its receiving stream. Environ. Toxi-col. Chem. 8, 437–450.

Biselli, S., Bester, K., Huhnerfuss, H., Fent, K., 2000. Concentrationsof the antifouling compound Irgarol 1051 and of organotins inwater and sediments of German North and Baltic Sea marinas.Mar. Pollut. Bull. 40, 233–243.

Brack, W., Altenburger, R., Ensenbach, U., Moder, M., Segner, H.,Schuurmann, G., 1999. Bioassay-directed identification of or-ganic toxicants in river sediment in the industrial region of Bit-

B inC-1

B e-ron.

B .S.,udk,

B as asys-

B as asys-

B ress,

C y ofh cell

Chau, Y.K., Maguire, R.J., Brown, M., Yang, F., Batchelor, S.P., 1997.Occurrence of organotin compounds in the Canadian aquatic en-vironment five years after the regulation of antifouling uses oftributyltin. Water Qual. Res. J. Can. 32, 453–521.

Clements, W.H., Kiffney, P.M., 1994. Integrated laboratory andfield approach for assessing impacts of heavy metals at theArkansas River, Colorado. Environ. Toxicol. Chem. 13, 397–404.

Clemons, J.H., Lee, L.E.J., Myers, C.R., Dixon, D.G., Bols,N.C., 1996. Cytochrome P4501A induction by polychlorinatedbiphenyls (PCBs) in liver cell lines from rat and trout and thederivation of toxic equivalency factors (TEFs). Can. J. Fish.Aquat. Sci. 53, 1177–1185.

Diamond, J., Daley, C., 2000. What is the relationship between wholeeffluent toxicity and instream biological condition? Environ. Tox-icol. Chem. 19, 158–168.

EC, 1994. Risk assessment of existing substances: technical guidancedocument, XI/919/94-EN No. European Commission.

EC, 1996. Technical Guidance Document in Support of CommissionDirective 93/67/EEC on Risk Assessment for New Notified Sub-stances and Commission Regulation (EC) No. 1488/94 on RiskAssessment for Existing Substances. European Commission.

ECETOC, 1993. Environmental hazard assessment of substances.ECETOC Document No. 51. European Centre for Ecotoxicologyand Toxicology of Chemicals.

Fent, K., 1995. Wirkungen von Spurenstoffen. EAWAG News 40D,8–12.

Fent, K., 1996. Ecotoxicology of organotin compounds. Crit. Rev.Toxicol. 26, 1–117.

Fent, K., 1998. Effects of organotin compounds in fish: from themolecular to the population level. In: Braunbeck, T., Hinton,D.E., Streit, B. (Eds.), Fish Ecotoxicology. Birkhauser Verlag,Basel, pp. 259–302.

Fent, K., 2001. Fish cell lines as versatile tools in ecotoxicology:assessment of cytotoxicity, cytochrome P4501A induction po-

ntal

FF ten-

omauc-

47–

F ta of

F ivers:icol.

F ntal42

F ty ofes.

F life.

terfeld. Arch. Environ. Contam. Toxicol. 37, 164–174.ruschweiler, B.J., Wurgler, F.E., Fent, K., 1995. Cytotoxicity

vitro of organotin compounds to fish hepatoma cells PLH(Poeciliopsis lucida). Aquat. Tox. 32, 143–160.

ruschweier, B.J., Wurgler, F., Fent, K., 1996. An ELISA for the dtermination of cytochrome P4501A in fish cell cultures. EnviToxicol. Chem. 15, 592–596.

ryan, G.W., Gibbs, P.E., Huggett, R.J., Curtis, L.A., Bailey, DDauer, D.M., 1989. Effects of tributyltin pollution on the msnail,Ilyanassa obsoleta, from the York River and Sarah CreeChesapeake Bay. Mar. Pollut. Bull. 20, 458–462.

ucheli, T.D., Fent, K., 1995. Induction of cytochrome P450biomarker for environmental contamination in aquatic ecotems. Crit. Rev. Environ. Sci. Technol. 25 (3), 201–268.

ucheli, T.D., Fent, K., 1996. Induction of cytochrome P450biomarker for environmental contamination in aquatic ecotems. Crit. Rev. Environ. Sci. Technol. 25, 201–268.

urton, G.A., 1992. Sediment Toxicity Assessment. CRC PBoca Raton, FL, USA.

astano, A., Vega, M.M., Tarazona, J.V., 1995. Acute toxicitselected metals and phenols on RTG-2 and CHSE-214 fislines. Bull. Environ. Contam. Toxicol. 55, 222–229.

tential and estrogenic activity of chemicals and environmesamples. Toxicol. In Vitro 15, 477–488.

ent, K., 2003.Okotoxikologie. Thieme-Verlag, Stuttgart.ent, K., Batscher, R., 2000. Cytochrome P450A induction po

cies of polycyclic aromatic hydrocarbons in a fish hepatcell line: demonstration of additive interactions and an indtion equivalency concept. Environ. Toxicol. Chem. 19, 202058.

ent, K., Hunn, J., 1991. Phenyltins in water, sediment, and biofreshwater marinas. Environ. Sci. Technol. 25, 956–963.

ent, K., Hunn, J., 1995. Organotins in freshwater harbors and rtemporal distribution, annual trend and fate. Environ. ToxChem. 14, 1123–1132.

ent, K., Hunn, J., 1996. Cytotoxicity of organic environmechemicals to fish liver cells (PLHC-1). Mar. Environ. Res.(1–4), 377–382.

ent, K., Looser, P.W., 1995. Bioaccumulation and bioavailabilitributyltin chloride: influence of pH and humic acids. Water R29, 1631–1637.

ent, K., Meier, W., 1992. Tributyltin-induced effects on earlystages of minnowsPhoxinus phoxinus. Arch. Environ. ContamToxicol. 22, 428–438.


Fent, K., Muller, M.D., 1991. Occurrence of organotins in municipalwastewater and sewage sludge and behavior in a treatment plant.Environ. Sci. Technol. 25, 489–493.

Fent, K., Stegeman, J.J., 1991. Effects of tributyltin chloride on thehepatic microsomal monooxygenase system in the fishStenoto-mus chrysops. Aquat. Toxicol. 20, 159–168.

Fent, K., Stegeman, J.J., 1993. Effects of tributyltin in vivo on hep-atic cytochrome P450 forms in marine fish. Aquat. Toxicol. 24,219–240.

Fent, K., Woodin, B.R., Stegeman, J.J., 1998. Effects of triphenyltinand other organotins on hepatic monooxygenase system in fish.Comp. Biochem. Physiol. C 121, 277–288.

Fioramonti, E., Semlitsch, R.D., Reyer, H.-U., Fent, K., 1997. Effectsof triphenyltin and pH on the growth and development ofRanalessonaeandRana esculentatadpoles. Environ. Toxicol. Chem.16, 1940–1947.

Guruge, K.S., Tanabe, S., Iwata, H., Taksukawa, R., Yamagishi,S., 1996. Distribution, biomagnification, and elimination ofbutyltin compound residues in common cormorants (Phalacro-corax carbo) from Lake Biwa, Japan. Arch. Environ. Contam.Toxicol. 31, 210–217.

Hahn, M.E., 1998. The aryl hydrocarbon receptor: A comparativeperspective. Comparative Biochemistry and Physiology. 121,23–53.

Hahn, M.E., Lamb, T.M., Schultz, M.E., Smolowitz, R.M., Stege-man, J.J., 1993. Cytochrome P4501A induction and inhibitionby 3,3′,4,4′-tetrachlorobiphenyl in an Ah receptor-containing fishhepatoma cell line (PLHC-1). Aquat Toxicol. 26, 185–208.

Hahn, M.E., Woodward, B.L., Stegeman, J.J., Kennedy, S.W., 1996.Rapid assessment of induced cytochrome P4501A protein andcatalytic activity in fish hepatoma cells grown in multiwell plates:response to TCDD, TCDF and two planar PCBs. Environ. Toxi-col. Chem. 15 (4), 582–591.

Hamasaki, T., Sato, T., Nagase, H., Kito, H., 1993. The mutagenicityof organotin compounds as environmental pollutants. Mutat. Res.

H an,r ex-etal-

H onBank

H cha-nol.

H altersydro-.

H cell.

H ctson

H .E.,mu-

lation in PLHC-1 fish cells exposed to sediment and oil shaleextracts. Arch. Environ. Contam. Toxicol. 38, 59–69.

Iwata, H., Tanabe, S., Mizuno, T., Tatsukawa, R., 1995. High accu-mulation of toxic butyltins in marine mammals from Japanesecoastal waters. Environ. Sci. Technol. 29, 2959–2962.

Jak, R.G., Ceulmans, M., Scholten, M.C.T., Van Straalen, N.M.,1998. Effects of tributyltin on a coastal North Sea plankton com-munity in enclosures. Environ. Toxicol. Chem. 17, 1840–1847.

Jung, D., Klaus, T., Fent, K., 2001. Induction of cytrochrome P450by nitroaromatic hydrocarbons and azaarens in the fish hepatomacell line PLHC-1. Environ. Toxicol. Chem. 20, 149–159.

Kannan, K., Corsolini, S., Focardi, S., Tanabe, S., Tatsukawa, R.,1996. Accumulation pattern of butyltin compounds in dolphin,tuna, and shark collected from Italian coastal waters. Arch. En-viron. Contam. Toxicol. 31, 19–23.

Kannan, K., Senthilkumar, K., Loganathan, B.G., Takahashi, S.,Odell, D.K., Tanabe, S., 1997. Elevated accumulation oftributyltin and its breakdown products in bottlenose dolphins(Tursiops truncatus) found stranded along the US Atlantic andGulf coasts. Environ. Sci. Technol. 31, 296–301.

Kannan, K., Tanabe, S., Iwata, H., Tatsukawa, R., 1995. Butyltins inmuscle and liver of fish collected from certain Asian and Ocea-nian countries. Environ. Pollut. 90, 279–290.

Kim, G.B., Tanabe, S., Tatsukawa, R., Loughlin, T.R., Shimazaki,K., 1996. Characteristics of butyltin accumulation and its bio-magnification in Steller sea lion (Eumetopias jubatus). Environ.Toxicol. Chem. 15, 2043–2048.

Koller, G., Hungerbuhler, K., Fent, K., 2000. Data ranges in aquatictoxicity of chemicals—consequences for environmental riskanalysis. Environ. Sci. Pollut. Res. 7, 135–143.

Linthurst, R.A., Bourdeau, P., Tardiff, R.G., 1995. Methods to Assessthe Effects of Chemicals on Ecosystems. Wiley, New York.

Looser, P.W., Bertschi, S., Fent, K., 1998. Bioconcentration andbioavailability of organotin compounds: influence of pH and hu-mic substances. Appl. Organomet. Chem. 12, 601–611.

L ach,ndss-4,

M iron.

M encevi-

M h-uscs38,

O role1.

O ima,o-in

P ucern-

300, 265–271.ansen, J.A., Woodward, D.F., Little, E.E., DeLonay, A.J., Bergm

H.L., 1999. Behavioural avoidance: possible mechanism foplaining abundance and distribution of trout species in a mimpacted river. Environ. Toxicol. Chem. 18, 313–317.

augg, M., Ackermann, G., Fent, K., 2000. Molecular cloningan estrogen receptor beta subtype from rainbow trout. Genaccession number AJ289883.1.

ermosin, M.C., Piedad, M., Cornejo, J., 1993. Adsorption menisms of monobutyltin in clay minerals. Environ. Sci. Tech27 (12), 2606–2611.

estermann, E.V., Stegeman, J.J., Hahn, M.E., 2000. Serumthe uptake and relative potencies of halogenated aromatic hcarbons in cell culture bioassays. Toxicol. Sci. 53, 316–325

ightower, L.E., Renfro, J.L., 1988. Recent applications of fishculture to biomedical research. J. Exp. Zool. 248, 290–302

origuchi, T., Shiraishi, H., Shimizu, M., Morita, M., 1997. Effeof triphenyltin chloride and five other organotin compoundsthe development of imposex in the rock shell,Thais clavigera.Environ. Pollut. 95, 85–91.

uuskonen, S.E., Tuvikene, A., Trapido, M., Fent, K., Hahn, M2000. Cytochrome P4501A induction and porphyrin accu

ooser, P.W., Fent, K., Berg, M., Goudsmit, G.H., SchwarzenbR.P., 2000. Uptake and eliminationof triorganotin compouby larval midgeChironomus ripariusin the absence and preence of Aldrich humic acid (AHA). Environ. Sci. Technol. 35165–5171.

ackay, D., 1982. Correlation of bioconcentration factors. EnvSci. Technol. 16, 274–278.

athiessen, P., Gibbs, P.E., 1998. Critical appraisal of the evidfor tributyltin-mediated endocrine disruption in mollusks. Enron. Toxicol. Chem. 17, 37–43.

orcillo, Y., Porte, C., 1997. Interaction of tributyl- and tripenyltin with the microsomal monooxygenase system in molland fish from the western Mediterranean. Aquat. Toxicol.35–46.

ehlmann, J., Bettin, C., 1994. TBT-induced imposex and theof steroids in marine snails. Malacol. Rev. Suppl. 6, 157–16

mura, M., Ogata, R., Kubo, K., Shimasaki, Y., Aou, S., OshY., Tanaka, A., Hirata, M., Makita, Y., Inoue, N., 2001. Twgeneration reproductive toxicity study of tributyltin chloridemale rats. Toxicol. Sci. 64, 224–232.

etrulis, J.R., Brunce, N.J., 1999. Competitive inhibition by indas a confounding factor in the use of the ethosyresorufio-


deethylase (EROD) assay to estimate exposure to dioxin-likecompounds. Toxicol. Lett. 105, 251–260.

Safe, H., 1994. Polychlorinated biphenyls (PCBs): environmentalimpact, biochemical and toxic responses, and implications forrisk assessment. Crit. Rev. Toxicol. 24 (2), 87–149.

Segner, H., 1998. Fish cell lines as a tool in aquatic toxicology.In: Braunbeck, T. (Ed.), Fish Ecotoxicology. Birkhauser Verlag,Basel, pp. 1–38.

Stab, J.A., Traas, T.P., Stroomberg, G., van Kesteren, J., Leonards,P., van Hattum, B., Brinkman, U.A.T., Cofino, W.P., 1996. De-termination of organotin compounds in the foodweb of a shallowfreshwater lake in the Netherlands. Arch. Environ. Contam. Tox-icol. 31, 319–328.

Stegeman, J.J., Hahn, M.E., 1994. Biochemistry and molecularbiology of monooxygenases: current perspectives on forms, func-tions and regulation of cytochrome P450 in aquatic species. In:Ostrander, G.K., Malins, D. (Eds.), Biochemical and MolecularBiological Perspectives in Aquatic Toxicology. Lewis Publish-ers, Boca Raton, pp. 87–206.

Takahashi, S., Tanabe, S., Takeuchi, I., Miyazaki, N., 1999. Dis-tribution and specific bioaccumulation of butyltin compoundsin a marine ecosystem. Arch. Environ. Contam. Toxicol. 37,50–61.

Tanabe, S., 1999. Butyltin contamination in marine mammals—areview. Mar. Pollut. Bull. 39, 62–72.

Traunspurger, W., Haizer, M., Hoss, S., Beier, S., Ahlf, W., Stein-berg, C., 1997. Ecotoxicological assessment of aquatic sediments

with Caenorhabditis elegans(nematoda)—a method for testingin liquid medium and whole sediment samples. Environ. Toxicol.Chem. 16, 245–250.

Unger, M.A., MacIntyre, W.G., Huggett, R.J., 1988. Sorption behav-ior of tributyltin on estuarine and freshwater sediments. Environ.Toxicol. Chem. 7, 907–915.

Villeneuve, D.L., Crunkilton, R.L., DeVita, W.M., 1997. Aryl hydro-carbon receptor-mediated toxic potency of dissolved lipophilicorganic contaminants collected from Lincoln Creek, Milwaukee,Wisconsin, USA, to PLHC-1 (Poeciliopsis licida) fish hepatomacells. Environ. Toxicol. Chem. 16, 977–984.

Vondracek, V., Machala, M., Minksova, K., Blaha, L., Murk,A.J., Kozubik, A., Hofmanova, J., Hilscherova, K., Ulrich, R.,Ciganek, M., Neca, J., Svrckova, D., Holoubek, I., 2001. Moni-toring river sediments contaminated predominantly with polyaro-matic hydrocarbons by chemical and in vitro bioassay techniques.Environ. Toxicol. Chem. 20, 1499–1506.

Willet, K.L., Gardinali, P.R., Sericano, J.L., Wade, T.L., Safe, S.H.,1997. Characterization of the H4IIE rat hepatoma cell bioassayfor evaluation of environmental samples containing polynucleararomatic hydrocarbons (PAHs). Arch. Environ. Contam. Toxicol.32, 442–448.

Zabel, E.W., Pollenz, R., Peterson, R.E., 1996. Relative potencies ofindividual polychlorinated dibenzo-p-dioxin, dibenzofuran, andbiphenyl congeners and congener mixtures based on inductionof cytochrome P4501A mRNA in a rainbow trout gonadal cellline (RTG-2). Environ. Toxicol. Chem. 15 (12), 2310–2318.

Documents

Ecotoxicological effects at contaminated sites