Xenobiotic Substances Such as PCB Mixtures(Aroclor 1254) and TBT Can InfluenceSwimming Behavior and BiotransformationActivity (GST) of Carp (Cyprinus carpio)

Katja Schmidt, Christian E. W. Steinberg, Stephan Pflugmacher, Georg B. O. Staaks

Leibniz-Institute of Freshwater Ecology and Inland Fisheries, Muggelseedamm 301,D-12587 Berlin, Germany

Received 11 August 2003; revised 23 February 2004; accepted 25 March 2004

ABSTRACT: Different groups of carp were treated with polychlorinated biphenyl (PCB) or tributyltin (TBT),and possible effects of the chemicals on the swimming behavior of the carp were examined using theBehavioQuant system. By evaluating quantitative behavioral parameters of the animals, it became evidentthat exposure to high concentrations of chemicals (organotin, 7 �g L�1, or polychlorinated biphenyl, 22�g L�1) severely affected the carp, causing a significant change in their swimming speed. TBT stress ledto a rapid decrease in mean swimming activity. A decrease in the preferred swimming depth wasobserved in TBT- and PCB-exposed fish. Animals exposed to PCB reduced their mean daily activity andincreased their mean swimming speed in the nighttime during the second week of exposure. Therefore,our findings imply that the fish were adapted to cope with the chemicals after the second week ofexposure. Furthermore, our results showed that low concentrations (TBT, 0.3 and 2 �g L�1, or PCB, 14�g L�1) did not significantly alter any quantified parameters of swimming behavior. In addition, the directeffects of chemicals on enzyme activity (GST) were determined. Measurement of soluble glutathione-S-transferase activity of fish liver or gills showed a significant elevation after exposure to PCB (0.1 or 22 �gL�1) or TBT (0.8 or 7 �g L�1). We had to conclude that the two different end points tested generally areuseful as biomarkers of exposure and for investigations of energy resources in organisms under theinfluence of toxic stress. © 2004 Wiley Periodicals, Inc. Environ Toxicol 19: 460–470, 2004.

Keywords: chronic toxicity; PCB; polychlorinated biphenyls; TBT; tributyltin, BehavioQuant; swimmingactivity; Cyprinus carpio; biotransformation; GST; glutathione-S-transferase

INTRODUCTION

As a result of pollutants being transported from industrialareas into the environment and their chemical persistence

once there, many freshwater ecosystems are faced withspatially or temporally alarming high levels of xenobiotics(Brack et al., 2002; Diez et al., 2002). The compoundsstudied in the present experiments were tributyltin (TBT)and polychlorinated biphenyl (PCB), which were selectedbecause they are representative of man-made chemicals inthe environment.

Both TBT and PCB are known to be toxic, and severaleffects have been reported for fish as well. TBT is widelyused as an additive in antifouling paintings, and therefore

Correspondence to: K. Schmidt; e-mail: schmidt@igb-berlin.de

Contract grant sponsor: Deutsche Bundesstiftung Umwelt, Germany.Contract grant number: 6000/708.

Published online in Wiley InterScience (www.interscience.wiley.com).DOI 10.1002/tox.20051

© 2004 Wiley Periodicals, Inc.

460

significant concentrations of TBT can be measured in nat-ural waters (5.76 �g L�1, Maguire et al., 1986, reported inGrinwis et al., 1998). Furthermore, exposure of fish toorganotin in concentrations on the same order of magnitudeas maximum TBT levels measured in the field caused mor-tality after 7–12 days resulted in gill lesions (17.3 �g L�1;Grinwis et al., 1998). Douglas et al. (1986) described the96-h LC50 for rainbow trout as 11.2 �g L�1. Dose-depen-dent teratogenic effects and delayed hatching also werereported (Fent, 1996), and TBT was shown to induce im-posex in female neogastropods of different species (Schulte-Oehlmann et al., 1996). In contrast to the androgenic effectsof TBT, PCBs are known as estrogen substances.

Mixtures of PCB isomers with different ingredients havebeen used in the mechanical engineering and the chemicaland electrotechnical industries. Restrictions on the use ofPCBs have resulted in considerable reductions in their con-centrations in surface water. However, PCBs are known tobe highly persistent, and therefore they occur in freshwatersediment at concentrations of, for example, 9.6–567 �g g�1

dw (Kannan et al., 1997). Moreover, significant quantitiesof PCBs are still detectable in most parts of the environment(Koponen et al., 2000). Because PCBs are lipophilic andhave the potential to accumulate, high levels of PCB weremeasured in aquatic animals, up to 2.2 �g PCB g�1 egg ww(Koponen et al., 2000). PCB are widespread environmentalpollutants that cause a broad spectrum of toxic effects invertebrates. For example, they can induce damage to theparenchymatous organs, mainly the liver, including hyper-trophy, local dystrophy, and necrobiotic to necrotic changes(Svobodova et al., 1994). In particular, Aroclor 1254 pro-duces cellular damage in the digestive gland in the form ofdecreased epithelial digestive cell height and decreased ly-sosomal membrane stability (Livingstone et al., 1997).

Most of the recent research on the influence of xenobi-otics on fish has described acute toxic effects (Douglas etal., 1986; Grinwis et al., 1998), sublethal and chronic influ-ences (Steinberg et al., 1995; Koponen et al., 2000), or hasfocused on in vitro experiments with cell lines (Pang et al.,1999; Brack et al., 2002). Even connections between fishdisease (M74 syndrome) and xenobiotic stress have beenstudied (Lundstrom et al., 1999). However, little on thebehavioral parameters of fish has been reported (Steinberget al., 1995; Spieser et al., 2001). In our opinion, thebiological function of behavior belongs to the reaction andadaptation of organisms to their environment (particularly,an uncomfortable, toxic environment), and therefore, quan-titative behavior analysis has proved to be an appropriatebioindicator for stress situations. Also, internal biochemicaladaptations such as the activation of biotransformation en-zyme systems in the presence of xenobiotic substancesenable the organism to survive in subacute toxic situations.

Much is known about the impact of both chemicals onenzyme activity (Safe, 1990; Fent, 1996; Otto and Moon,1996; Oberdorster et al., 1998; Whyte et al., 2000; Perez-

Lopez et al., 2002). PCB is a potent inducer of enzymesystems such as the microsomal CYP1A monooxygenasesystem (EROD) and the conjugation enzyme system (solu-ble glutathione-S-transferase) in fish (Koponen et al., 2000).In contrast to these findings, the EROD activity of TBT-exposed crabs showed that CYP1A was not elevated (Ob-erdorster et al., 1998), but Morcillo and Porte (1997) de-scribed a clear inhibition of EROD activity of themicrosomal fraction isolated from the fish liver. However,fish are able to both behaviorally and physiologically adaptto xenobiotics.

Our experiments focused on the observed sublethal ef-fects of TBT or PCB on animals. The first objective of ourstudy was to determine the daily pattern of swimmingbehavior and to describe the normal synchronized rhythmsof this parameter. We investigated whether carp changes itsbehavior (e.g., swimming activity or preferred areas) in anuncomfortable environment because of chemical stress, andin addition, we measured the activity of the detoxicationenzyme system (GST) after exposure to both xenobiotics. Inconclusion, the present article discusses the interaction be-tween the available energy for locomotion and biotransfor-mation in the presence of foreign substances.

MATERIALS AND METHODS

Cyprinus carpio and Rearing Conditions

The carp used were obtained as fry from the aquaculturedepartment of our institute (IGB Berlin, Germany). Theanimals were held in several 80-L aquariums supplied withtap water at approximately 24°C � 1°C under a 12 h:12 hphotoperiod. The animals were fed twice daily with eitherdry or frozen mixed food (Tetramin, Daphnia, red Chirono-midae larvae). The Tetramin used was a commercial dietcontaining approximately 46% protein and 6% lipids. Onlyfish that weighed 1.4 � 0.3 g and had a length of 46 � 2mm were used in the experiments.

Chemicals and Chemical Analysis

We used two different xenobiotic substances: a technicalPCB mixture (Aroclor 1254, Boeringer, Ingelheim, Ger-many) or tributyltin chloride (TBT; Sigma-Aldrich,Schnelldorf, Germany). The PCB purity was 98%, and theTBT purity was 96%. At the beginning of each experimentalseries, concentrated dilutions of TBT or PCB were preparedwith solubilizer methanol (purity of 99.9%) and stored at4°C. Portions of these stock solutions were diluted withdistilled water daily to obtain the final fish test concentra-tion. To measure the actual aqueous concentrations of PCBor TBT, fish medium was collected from each experimentduring the exposition period; the chemical analysis of thesewater samples was carried out by means of GC/MS (or-

XENOBIOTICS INFLUENCE CARP SWIMMING/GST 461

ganotin; DIN 38407-13) and GC/ECD (PCB; DIN 38407-3). The concentrations used in the experiments were:

PCB: 0.1, 14, and 22 �g L�1;TBT: 0.3, 0.8, 2, and 7 �g L�1.

BehavioQuant System

The swimming behavior of individual carp was recorded toassess sublethal effects in an uncomfortable environment byusing the automated video analysis system BehavioQuant(Metacom, Germany).

Test aquariums were continuously observed by videocameras, and the positions of each single animal wererecorded according to Steinberg et al. (1995) and Baganz etal. (1998). Fish measured simultaneously in each aquariumwere recognized as dark images on an illuminated back-ground. The pictures of video cameras were analyzed at aspeed of 24 frames per second. Each of the six aquariumswas studied for 2-min time periods at 20-min intervals over23 h from 10:00 a.m. to 9:00 a.m. the following day, making68 observation intervals per day. The images were digitized,and the 2-min results were averaged into one readout. Sub-sequently, the swimming parameters of animals were ana-lyzed: motility (swimming speed in pix s�1) and swimmingdepth (below surface in pix).

Experimental Conditions

Fish tests were approved by Landesamt fur Arbeitsschutz,Gesundheitsschutz und technische Sicherheit (Berlin, Ger-many; license no. G0292/00). Experiments were carried outin a flow-through system with 24-L test chambers. Theswimming space of each aquarium measured 40 � 30 � 30cm (length � depth � width). The chemical dilutions (TBTor PCB) were added to stock aquariums with continuallyaerated water at the predefined concentrations. From therethe solution renewal for each aquarium was 10 L per day.Inside each aquarium the water was additionally aerated byrecirculation. The resulting slow flow rate inside did notstress or disturb the fish.

Shoals of six fish each were picked from the stockcultures and transferred to the test aquariums after measure-ment of body weight and length. After a period of acclima-tization, which last approximately 1 week, video recordingwas started. Then the fish were observed for 3 weeks duringwhich they were subject to any chemical stress (21-daynonexposure period); thereafter, they were exposed to thexenobiotic substances (21-day exposure period). The exper-iments were terminated after 21 days of exposure, or earlierif fish showed obvious pathologic effects. Some groupsserved as controls, and the different exposures were exam-ined in a minimum of two replicates.

A temperature of 20°C and a 12 h:12 h light–dark cycle(from 8:00 a.m. to 8:00 p.m.) using 60 W halogen- and

infrared- (permanently) lamps were maintained constantly.Fish were fed the equivalent of approximately 3% of theirbody mass per day. Animals got a frozen-food mixture(50% Daphnia and 50% red Chironomidae larvae) at 10:00p.m.; at 3:00 p.m. automatic feeders delivered a dry foodmixture (45% Tetramin, 30% Daphnia, and 25% Chirono-midae). From 9:00 p.m. to 10:00 p.m. each aquarium wascleaned mechanically, and waste was removed.

The variation between measurements with and withoutsolubilizer used for the stock solutions was in no casestatistically significant, and exposure conditions were basi-cally the same as those described for the xenobiotics treat-ments.

Enzyme Measurement

Enzymes of the fish liver or gills were prepared according toPflugmacher and Steinberg (1997) with modifications listedbelow by homogenization, centrifugation, saturation, anddesalting. Samples were homogenized in sodium phosphatebuffer (NaP, pH 6.5, 0.1 M) with glycerol (20%) dithio-erythriol (DTE 1.4 mM) and ethylenediaminetetraaceticacid (EDTA; 1 mM). The material was centrifuged at6500 � g for 10 min, and the resulting supernatant wascentrifuged again at 104 000 � g for 1 h. After the firstsaturation of the supernatant, defined as the soluble fraction,with ammonium sulfate (35%) followed by centrifugation at26 000 � g for 20 min, the pellet was discharged. Theresulting supernatant was used for second precipitation(80%), followed by centrifugation at 58 500 � g for 30 min.The pellet was resuspended in sodium phosphate buffer(NaP, pH 7, 20 mM) with DTE (0.6 mM) and desalted usingan NAP.10 column.

After this extraction procedure, the material was frozenin liquid nitrogen and stored at �80°C until measurement ofspecific enzyme activity.

The enzyme activity of soluble glutathione-S-transferase(GST) was determined by a spectrophotometer according toHabig et al. (1974) using 1-chloro-2,4-dinitrobenzene(CDNB) as standard model substrate. The enzyme activityof the GST was measured at 340 nm and 30°C during therecording time of 5 min. Finally, the protein content of thesamples was measured according to Bradford (1976). En-zyme activity is given in nkat mg�1 protein.

Statistical Comparison

Behavioral data are expressed as mean � standard deviationfor each measuring period of 2 min. They were calculated asmean � standard deviation for 1 h or for day and nightperiods. The statistical differences of behavioral variableswere assessed using the Wilcoxon rank test and the Mann–Whitney U test. The differences between enzyme systemactivities were analyzed using the Mann–Whitney U test

462 SCHMIDT ET AL.

(Siegel, 1985). All statistical analyses were calculated withn indicating the number of animals examined and withsignificance determined at the 95% level(p � 0.05).

RESULTS

Swimming Behavior of C. carpio inUnexposed Conditions

Several experiments were done with shoals of six animals toassess the swimming behavior of carp without any chemicalstress. No mortality occurred in any of these experiments.

Fish rapidly acclimated to experimental conditions. Theswimming activity of fish of the same shoal was found to becomparable on all days. Figure 1(A) displays the meanmotility (of six different shoals) under comfortable condi-tions. The swimming activity registered was circadian, syn-chronized with time; local activity peaks and reduced ac-tivity periods were observed in all groups. The rhythmobserved was synchronized to food availability (10:00 a.m.and 3:00 p.m.) and light switching (8:00 a.m. and 8:00p.m.). Fish preferred mainly lower layers near the bottom ofthe aquarium during the light period (except during feeding

periods) but preferred a mean swimming depth nearer to thesurface during darkness [Fig. 1(B)].

It was presumed that the results of this experimentalseries were be the normal swimming behavior of carp overthe whole measuring period, which did not differ betweengroups.

Swimming Behavior of C. carpio Exposed toPCB

No mortality occurred during either the nonexposure period(animals observed under comfortable conditions) or theexposure period (fish contaminated with PCB). In contrastto the high-level-exposed (22 �g L�1) animals, the behav-ioral parameters of fish exposed to the low concentration(14 �g L�1) did not differ from the controls. Circadianorganization of swimming behavior of low-level-exposedcarp and the control fish did not vary over the wholemeasuring time.

The results of high-level exposure compared to the con-trol are given in Figure 2. The behavioral parameters duringthe nonexposure and exposure periods are presented as dailymeans over the light and dark phases. We found significantchanges in both swimming speed and swimming depth. Thebehavior of carp when not exposed showed oscillations ofmotility, with high day values and low levels at night,simultaneously with a preferred low horizontal positionduring the day instead of the upper layers, which werepreferred at night.

It should be pointed out that, compared with the nontoxictime, the high-level test concentration (Fig. 2, from day 8 today 28) did not significantly affect the behavior during thefirst week. The following week motility was significantlyincreased during the night [p � 0.02; Fig. 2(A), 2, from day16 to day 22], which was clearly related to a dramaticallysignificant decrease in swimming depth at night [p � 0.02;Fig. 2(B), 2, from day 16 to day 18]. Daily means alsochanged significantly during the second week of exposure:both motility (p � 0.05) and swimming depth (p � 0.01)decreased. The processes during the second week of expo-sure were combined with an increasing number of fish,which showed comparatively poor motivation to feed. Aftersome days the animals became acclimated to the toxicexposure, and the second reaction, which was a spontaneousphase shift to an obviously normal swimming behavior (likethat under nontoxic conditions), followed.

In conclusion, these experiments demonstrated changesin the swimming behavior of carp as a result of a high levelof exposure to PCB (LOEC � 22 �g L�1). The xenobioticsubstance affected the motility and the swimming depth offish from the 9th to the 13th days of exposure. Thesebehavioral changes were reversible.

Fig. 1. Mean swimming behavior with standard deviationsof carp under comfortable conditions over 23 h (6 shoals).The swimming activity registered was circadian, synchro-nized with time, and the rhythm was synchronized to foodavailability (10:00 a.m. and 3:00 p.m.) and light regime (8:00a.m. and 8:00 p.m.). (A) Mean motility: carp swimming speedwas low at night and relatively high during the day; fish wereespecially active at feeding time. (B) Mean swimming depth:fish preferred lower layers near the bottom of the aquariumwhen it was light (except during feeding time), a swimmingdepth at the surface when it was dark.

XENOBIOTICS INFLUENCE CARP SWIMMING/GST 463

Swimming Behavior of C. carpio Exposed toTBT

After the nonexposure period the animals were exposed todifferent concentrations of TBT (0.3, 2, or 7 �g L�1). Nomortality occurred in these experimental series.

There were no significant differences in circadian swim-ming activity or preferred horizontal positions resultingfrom the lower concentration of the medium. The onlyeffects on behavioral variables that we found were at thehighest concentration of the medium.

Figure 3(B) displays the motility of one shoal exposed toa high-concentration medium (7 �g L�1) compared to thecontrol [Fig. 3(A)]. The nonexposure period showed a nor-mal graph, with swimming activity observed to be high atfeeding time and during light on/off switching, and motilityto be low during darkness. Animals did not change theirswimming speed over the preexposure time, but a greatimpact from TBT on fish behavior was observed from thefirst day of exposure on. Fish exposed to a high concentra-tion in the second part of the experiment [Fig. 3(B), fromday 4 to day 6] showed decreasing mean motility combined

with destroyed rhythms of swimming activity. We alsoobserved a significantly increased swimming speed in dark-ness (p � 0.01), whereas the fish preferred a lower level ofactivity during the daytime (p � 0.02). These findings werenoted to be nonreversible. The test was terminated on thefourth day of exposure because the carp were showingpathological symptoms (increased frequency of opercularbeats, disturbance of equilibrium, appetite depression). Un-eaten food was observed only in chambers in which fishwere exposed to a high level of TBT.

We estimated the LOEC to be 7 �g L�1 because thepresence of a high level of TBT and changes in swimmingbehavior were causally related, in contrast to the situation inlow-level-exposed animals.

Enzyme Activity of C. carpio Exposed to PCBor TBT

Basic activity of soluble glutathione-S-transferase (GST)was found in all fish tested including the control group, andactivity in the liver was always much higher than in the fish

Fig. 2. Mean swimming behavior in the daytime and in darkness of a shoal of six fish exposedto PCB compared to an unexposed shoal (control group). The preexposure period, days 1–7,was followed by the exposure period (PCB: 22 �g L�1), days 8–28. (A) 1: Mean motility ofcontrol group, 2: mean motility of PCB-exposed group. Animals exposed to PCB showedreduced mean daily activity and increased mean swimming speed in the nighttime only duringabout the second week of exposure. (B) 1: Mean swimming depth of control group; 2: meanswimming depth of PCB-exposed group. A decrease in the mean preferred swimming depthin PCB-exposed fish was observed only during days 16–18.

464 SCHMIDT ET AL.

gills. In general, measurement of the activity of the enzymesystem in carp after exposure to PCB or TBT showed directeffects on GST (the results are given in Figs. 4 and 5). Wefound an elevation of GST activity in the TBT group as wellas in the PCB-exposed fish, but the increased activity wasnot clearly dose related.

Biotransformation activity of fish liver as well as of gillswas clearly elevated (p � 0.01) in the PCB experiments (0.1or 22 �g L�1). The highest specific activity—elevation by91% compared to the controls—was found in fish liverexposed to PCB (0.1 �g L�1) with 40.9 � 6.4 nkat mg�1

protein. The corresponding value for the gills was 10.0 �2.2 nkat mg�1 protein. The higher PCB concentration (22�g L�1) caused a clear increase in GST activity in liver(35.1 � 6.9 nkat mg�1 protein) compared to the control, butGST activity (4.2 � 1.8 nkat mg�1 protein) in the gills ofthe high-level-PCB-exposed carp was not affected. Wefound no significant differences between the correspondingPCB treatments.

However, exposure to TBT (0.8 �g L�1) led to a signif-icant (p � 0.01) elevation (89%) of GST activity only ingills (10.4�3.1 nkat mg�1 protein) compared to the control(5.5 � 0.9 nkat mg�1 protein), but there was no significanteffect of TBT on enzyme activity in the carp liver with22.2 � 3.9 nkat mg�1 protein. Also, we found no significanteffects of a higher concentration of TBT (7 �g L�1) on GSTactivity in gills or liver after the third day of exposurecompared to the control.

DISCUSSION

Various biotests with fish have been used as instruments forthe estimation of acute and sublethal effects of chemicals inthe aquatic environment. Most of the recent research on theinfluence of toxins on animals has focused on species-dependent mortality, growth inhibition, or reproduction re-duction. Stress-dependent changes in behavior (Steinberg etal., 1995; Baganz et al., 1998; Spieser, 2000) or enzymeactivity (Pflugmacher and Steinberg, 1997; Koponen et al.,2000; Wiegand et al., 2000), also described in our study,also could act as an important biomarker for environmentalchemicals. Carp was chosen as the test organism in thecurrent study because it occupies a central position in fresh-water food webs, especially in inland fishery pond systems,which often are situated near industrial areas and are influ-enced by various xenobiotics. We tested the hypothesis thatPCB or TBT affects the swimming behavior and biotrans-formation enzyme system activity of carp.

Fig. 3. Mean swimming activity of a shoal of six fish (A)without exposure to TBT and (B) with exposure to TBT. Thepreexposure period, from days 1–3, was followed by theexposure period, (TBT: 7 �g L�1) during days 3 and 4. TBTled to a rapid decrease in daily swimming speed during theday and to an increased activity level at night.

Fig. 4. Activity of soluble glutathione-S-transferase in carpafter 21 days of exposure to PCB (0.1 or 22 �g L�1) in (A)liver and (B) gills. Data are means � standard deviations ofsix fish; a significant result (p � 0.05/U test) is marked withan asterisk.

Fig. 5. Activity of soluble glutathione-S-transferase in carpafter 21 days of exposure to TBT (0.8 or 7 �g L�1) in (A) liverand (B) gills. Tests with the highest concentration wereterminated after the third day of exposure (marked with across). Data are means � standard deviations of six fish; asignificant result (p � 0.05/U test) is marked with an asterisk.

XENOBIOTICS INFLUENCE CARP SWIMMING/GST 465

TBT and PCB are very toxic to fish. The LT50 of or-ganotins were described as 11.2 �g L�1 (96 h; Douglas etal., 1986) or 18 �g L�1 (14 days; Grinwis et al., 1998). Thefindings in the behavioral tests lie in the same order ofmagnitude. We measured a higher LOEC in the behavioralstudies (7 �g TBT L�1 and 22 �g PCB L�1) than in theenzyme tests (0.8 �g TBT L�1 and 0.1 �g PCB L�1). Byevaluating the quantitative behavioral parameters of ani-mals, it becomes evident that exposure to a high concentra-tion of a chemical medium (organotin, 7 �g L�1, or poly-chlorinated biphenyl, 22 �g L�1) significantly affects fish.A low concentration of a medium (TBT, 0.3 and 2 �g L�1,or PCB, 14 �g L�1) did not significantly alter any quantifiedswimming behavior. But a high level of exposure of carp tosingle chemicals caused a change in swimming activity:TBT stress led to a decrease in swimming speed. Animalsexposed to PCB reduced their mean daily activity andincreased their mean swimming speed in the night duringabout the second week of exposure, and after this phenom-enon the fish acclimated to the toxic influence. Knowing theincreased activity of the biotransformation enzyme in PCB-exposed fish, we concluded that the metabolic pathway ofthe xenobiotic enabled fish to adapt to the toxic influence.Possibly, therefore, the altered behavior during the secondweek of exposure was reversible. However, behavior isalways closely related to the ecological adaptation of organ-isms to their environment (Spieser, 2001). Even in toxicconditions fish have to be successful in the struggle for life.

Because there is evidence that animal activity dependson group size and social interactions between individuals,we always ran our behavioral experiments on shoals of sixfish. It generally has been shown that group formationsenable social facilitation of behavior (Warthold, 1980; Gril-litsch and Vogl, 1996; Boilliet et al., 2001). The respectiveimportance of photoperiod and feeding in the synchroniza-tion of swimming activity of fish was noted in all ourexperimental series. This is in agreement with the results ofseveral studies (Steinberg et al., 1995; Staaks, 1996; Baganzet al., 1998; Kamjunke et al., 2002) and suggests that in anatural environment fluctuations in the daily availability offood (e.g., diurnal vertical migration of fish food like Daph-nia species) or light conditions throughout the year might beone of the factors inducing diurnal or seasonal changes inswimming behavior. In this way the fish become spatiallyand temporally synchronized with the availability of naturalresources. Therefore, a phase shift in swimming speed or inpreferred layers in water space as a result of xenobioticstress (PCB or TBT), as reported in our study, may havegreat ecological relevance.

The BehavioQuant system used in our experiments alsocould be used as a biomonitoring system. However, unfor-tunately we found swimming behavior to be affected only athigh concentrations. Locomotion activity of a varietyaquatic organisms has been found to be changed by severalsubstances, even at very low concentrations (Steinberg et

al., 1995; Baganz et al., 1998). Effects of the herbicideatrazine on the swimming behavior of fish were studied bySteinberg et al. (1995). They reported that fish exposed toatrazine preferred dark areas of the aquariums. Anotherstudy found that the exposure of fish to the pure cyanobac-terial toxin microcystin had a dose–response effect on lo-comotion activity: whereas the two lower exposure concen-trations caused an increase in daytime motility, elevatedexposure concentrations led to decreased activity (Baganz etal., 1998). Moreover, reduced feeding and spawning activitywas noted at the highest test concentration of the toxin.

Also, in the present study low feeding motivation ofhighly exposed fish was associated with uneaten food par-ticles, whereas almost no food waste was observed in aquar-iums containing fish that experienced low or medium ex-posure. These fish also showed typical swimming behavior.The phenomenon of altered feeding behavior was observedin the PCB group during the second week of exposure andin the TBT-exposed fish much earlier (the third day). Thedecreased feed intake of the latter group could easily ex-plain the lower swimming speed as an energy-related re-sponse. In contrast, we did not find reduced motility of thePCB-exposed carp during reduced feeding (activity de-creased during the day and increased at night). Therefore,we concluded that carp reduced their energy expenditure fordigestion and put a higher amount of energy into biotrans-formation processes.

Depression of appetite is a common response of fish tostress, and intermittent feeding for long periods can have alarge effect on growth (Rice, 1990). In our study the in-crease in body growth in the TBT-exposed group was19.6% less than in the control group. This result is consis-tent with the findings of a study by Hartl et al. (2001), whodescribed only a minimal increase in body length of TBT-exposed fish. There seem to be enhanced metabolic costsattached to the changes caused by TBT exposure, which aremanifested as less increase in body length (Hartl et al.,2001): the control group showed a 7%–12% increase inlength, whereas the TBT group only increased in length by2%.

In our experiments with TBT and PCB, only the TBT-exposed shoals of carp showed decreased growth. Thesefindings are in accord with those of Jorgenson et al. (1999),who reported that PCB exposure (Aroclor 1264) did notaffect either growth or organ lipid concentration. The lowfeeding motivation of the highly exposed fish in our PCBgroup was noted only temporarily; during about the secondweek of exposure uneaten food particles were removed. Atthe end of the experiment no differences between the expo-sure phase and the nonexposure time were observed, and weconcluded that the PCB-exposed fish were well adapted tothe xenobiotic.

The substantial growth reduction caused by environmen-tal stress has implications for both aquaculture productionresults and survival under natural conditions. The effects of

466 SCHMIDT ET AL.

environmental stress at one trophic level may directly orindirectly influence other trophic levels, too; changes inswimming activity or feeding of individuals of one specieswill affect the survival, growth, or behavior of their preda-tors and prey. Carp in a natural environment must continu-ously expend energy for feeding activity and for predationavoidance. In toxic environmental conditions it might beprofitable for these animals to decrease either their swim-ming activity or their food uptake in order to lower theenergy costs of movement and digestion. On the other hand,small carp are common prey items for many piscivores.Individual size and behavioral abilities are important com-ponents of predator avoidance; thus, reduction in motilitymay change their survival rate dramatically.

The behavioral effect of xenobiotics on carp might be aresult of changes in physiologic-energy homeostasis. Ourhypothesis is that inhibited swimming activity (musclemovement) may be caused either by lowered general energystatus or by higher energy demand of other physiologicalprocesses. If the first explanation is valid, then the results ofour behavioral experiments may indicate that the observeddecrease in motility was only one sign of a general decreasein physiological activity induced by exposure to stress. Ifthe second explanation is valid, it means that sublethallyexposed fish should share the available energy betweenvarious physiological functions, for example, growth,swimming activity, and defense against toxic compounds.The latter process occurs in three phases (transformation,conjugation, and excretion) involving the induction of thedetoxication enzyme systems (Pflugmacher and Steinberg,1997): an organism can metabolize xenobiotics mainly byoxidation, reduction, hydrolysis (phase 1, catalyzed by, e.g.,CYT P450 monooxygenase), and conjugation (phase 2, e.g.,via glutathione-S-transferase) reactions. It seems that thisbiotransformation pathway is also valid for the PCB andTBT used in our experiments because our results showed asignificant elevation of the specific activity of the solubleGST after exposure to both PCB and TBT.

The enzyme system of phase 1 (cytochrome P450 mono-oxygenase system) was analyzed, and it was noted thatAroclor 1254 significantly induced hepatic EROD(ethoxyresorufin O-deethylase) activity (Brumley et al.,1995; Livingstone et al., 1997; Arukwe et al., 2000; Kopo-nen et al., 2000). Koponen et al. (2000) found that themicrosomal CYP1A monooxygenase system (EROD) was avery sensitive indicator of fish embryo exposure, beinginduced at low PCB concentrations even after only 3 days.By activating the Ah receptor, PCB may cause an inductionof cytochrome P4501A in fish, resulting in de novo proteinsynthesis. The principle reaction of the phase 2 enzymesystem (GST) was the conjugation of electrophilic sub-stances or groups to tripeptide glutathione [N-(N-L-glutanyl-L-cysteinyl)glycine] in order to make the xenobiotic morehydrophilic for transportation or excretion (Egaas et al.,1993; Pflugmacher et al., 2001). In contrast to our results,

Koponen et al. (2000) noted the absence of a clear PCB-exposure response in GST enzyme activities. Huuskonen etal. (1996) described a 75% increase in GST activity in liver,but they found that PCB did not affect conjugation activitiesas effectively as monooxygenase activities. Other authors(Otto and Moon, 1996; Perez-Lopez et al., 2002) also foundhigher GST activity in PCB-exposed fish than in the con-trols. Liver GST activities were threefold higher, whereas asimilar trend also was noted in kidney and white muscle,with activity 1.5 and 1.2 times higher (Otto and Moon,1996). This and the results of our experiments (91% in-creased activity of soluble GST in liver) support the hy-pothesis that the well-prepared (because of the hydroxylyticreaction catalyzed by the monooxygenase system) hydroxylgroup of the PCB substrate was conjugated by GST. Thisresult is in accord with the results presented by Perez-Lopezet al. (2002), who investigated GST in various fish organs(liver, kidney, gills). They also reported that liver samplesshowed the highest GST induction process, mainly associ-ated with a �-class-related isoenzyme. Also, a significantlylower total GSH level in exposed fish supports the involve-ment of GSH in the biotransformation process; furthermore,as GSH may be used by conjugation processes, the resultinglow GSH level may impair other metabolic mechanisms andthus weaken the physiological status of the organism (Ottoand Moon, 1996). Further metabolism of the resulting con-jugate could lead to a more water-soluble product for ex-cretion. This idea is supported by the increased activity ofuridine diphosphoglucoronosyltransferase (UDP-GT) foundby Brumley et al. (1995) and Huuskonen et al. (1996): thegreatest increases were in the liver (67%) and the kidney(241%). They concluded that metabolized PCB excretedinto the bile were hydroxylated and conjugated with glucu-ronic acid or sulfate.

The TBT (0.8 �g L�1) exposure in our experiments ledto the largest increase (89%) in soluble GST in fish gills,indicating the phase 2 biotransformation pathway of or-ganotin compounds. Therefore, TBT seems to be metabo-lized by conjugation catalyzed by GST in gills. The lack ofsignificant effects of higher TBT concentration (7 �g L�1)on the GST activity of liver or gills could be explained bythe short exposure time (3 days). There also was no statis-tically significant effect of the TBT on the liver enzymesystem in our exposure experiments, indicating that thephysiological limit of the GST might have been reached orthat the biotransformation enzyme was broken down. Onlybiotransformation activity in the gills (the tissue with directcontact to the xenobiotic medium) was shown to be sensi-tive enough to express GST activity after exposure to TBT,which also could be attributed to, for instance, the watersolubility of TBT in contrast to that of PCB. The inhibitioneffects of TBT on the phase 1 system were described byOberdorster et al. (1998) as in contrast to the clearly in-creased activity of GST (phase 2): TBT broke down cyto-chrome P450 enzymes in vitro. To determine in vivo effects,

XENOBIOTICS INFLUENCE CARP SWIMMING/GST 467

the P450 levels of crabs fed TBT were examined, and theEROD activity showed that CYP1A was not elevated.Therefore, Oberdorster et al. (1998) concluded that animalsup-regulate P450 isoenzymes, possibly to support the me-tabolism and elimination of TBT.

There were no clear differences in our study in theresponses seen in liver or gill GST activities between low-or high-level exposures. In the TBT group as well as inPCB-exposed fish, we found the highest enzyme elevationin the lower concentrations of the medium. Factors affectingthis lack of dose dependence may be a short exposure time(the TBT experiment was terminated after the third day) orthe enzyme system having already reached the functionalmaximum (PCB). In relation to the absence of a cleardose–response effect of PCB on GST activity found in theexperiments of Koponen et al. (2000), our results demon-strates that more research should be directed toward im-proving our understanding of the behavior of chemicals inorganisms.

In our opinion, all disturbances of the physiologicalsteady state of animals can induce adaptive mechanisms oreven pathological effects, which are energy demanding.Visual observation and the results of our laboratory testsindicated that the decrease in motility combined with theelevation of biotransformation activity is one part of gen-erally changed physiological activity as a result of toxicstress. Therefore, we calculated a Pearson correlation coef-ficient to see if there were a linear association betweenbiotransformation activity and swimming behavior. Weused data of the different concentrations at one time point(21 days) and found that the results of our TBT experimentsdescribed both a positive relation between preferred swim-ming height during the day and the enzyme activity in theliver samples (correlation coefficient r � 0.67) and a neg-ative relation between motility during the night and GSTactivity in gills (correlation coefficient r � �0.72). Parti-tioning of energy resources is one oft the most difficult tasksan organism has to solve under the influence of stress. Sothe energy spent for swimming can serve as an indicator ofthe costs of various other physiological processes, for in-stance, digestion and the biochemical processing of food orbiotransformation.

The change of habitat preferences from lower to upperlayers in our test aquariums during the second week of PCBexposure can be interpreted as a hint that the fish demandeda higher level of oxygen during the exposure period. If gillswere to be destroyed because of xenobiotics (Grinwis et al.,1998) or the membrane functions disturbed by changedpermeability (Hartl et al., 2001), oxygen uptake rate wouldrapidly decrease. Higher oxygen levels in surface areaswould enable the fish to compensate for their (possiblyhigher) oxygen demands and to maintain their global energystatus. As a consequence, the metabolic rate (in relationshipto the respiration) of fish could be increased under chemicalstress. In our opinion, the factor regulating the distribution

of energy between different physiological processes shouldbe investigated. Therefore, our further studies will investi-gate oxygen consumption of exposed fish combined withbehavioral measurements so that the cost of swimming andbiotransformation can be estimated.

CONCLUSIONS

In summary, the results presented indicate that the dose-dependent effects on fish swimming behavior and biotrans-formation enzyme activity were caused by PCB or TBTexposure. In our tests only high concentrations of chemicalsled to a decrease in swimming speed, an increase in swim-ming height, or phase shifting of activity. Even lower con-centrations of xenobiotics led to an elevation of detoxicationenzyme activity. The behavioral and physiological changesobserved are important toxicological end points showingsublethal disturbances of organism physiology and allowingdiscussion of ecological consequences. Carp must continu-ously expend energy for such activities as locomotion, andfor fish under sublethal toxic conditions, it might be prof-itable for them to decrease their swimming speed if theenergy input is lower than the cost of movement. Moreover,we pursed the hypothesis that the higher physiological en-ergy demand caused by increased activity of enzymes in-volved in the biotransformation of foreign substances maybe related to the higher respiration of fish stressed byxenobiotics. From our behavioral observations and biotrans-formation results, we could not differentiate whether theobserved changes in ecophysiological parameters weresymptoms of a partial energy shift or whether they wereonly signs of a global energy-saving strategy of fish or both.The partitioning of energy resources, for example, for de-toxication pathways, is one of the most difficult tasks anorganism has to solve under the influence of stress. Nodoubt, one possible result of destroyed physiological ho-meostasis of fish is a change in oxygen consumption. There-fore, respiration of fish should be investigated in furthertoxicological experiments in order to calculate the appor-tioning of energy in this organism.

The authors would like to acknowledge to Dr. K. Kohlmannand M. Kunow (IGB Berlin, Germany) for the breeding of carps.

REFERENCES

Arukwe A, Celius T, Walther B, Gogsoyr A. 2000. Effects ofxenoestrogen treatment on zona radiata protein and vitellogeninexpression in Atlantic salmon (Salmo salar). Aquat Toxicol49:159–170.

Baganz D, Staaks G, Steinberg C. 1998. Impact of the cyanobac-teria toxin microcystin-LR on behaviour of the zebrafish, Daniorerio. Water Res 32:948–952.

468 SCHMIDT ET AL.

Boilliet V, Aranda A, Boujjard T. 2001. Demand-feeding rhythmin rainbow trout and European catfish synchronisation by pho-toperiod and food availability. Physiol Behav 73:625–633.

Brack W, Schirmer K, Kind T, Schrader S, Schuurmann G. 2002.Effect-directed fractionation and identification of cytochromeP450A-inducing halogenated aromatic hydrocarbons in a con-taminated sediment. Environ Toxicol Chem 21:2654–2662.

Bradford MM. 1976. A rapid and sensitive method for the quan-tification of microgram quantities of protein utilizing the prin-ciple of protein-dye binding. Anal Biochem 72:248–254.

Brumley CM, Haritos VS, Ahokes JT, Holdway DA. 1995. Vali-dation of biomarkers of marine pollution exposure in sandflathead using Aroclor 1254. Aquat Toxicol 31:249–262.

Diez S, Abalos M, Bayona JM. 2002. Organotin contamination insediments from the western Mediterranean enclosures following10 years of TBT regulation. Water Res 36:905–918.

Douglas MT, Chanter DO, Pell IB, Burney GM. 1986. A proposalfor the reduction of animal numbers required for the acutetoxicity for fish test (LC50 determination). Aquat Toxicol 8:243–249.

Egaas E, Skaare JU, Svendsen NO, Sandvik M, Falls JG, Dauter-man WC, Collier TK, Netland J. 1993. A comparative study ofeffects of atrazine on xenobiotic metabolizing enzymes in fishand insect, and of the in vitro phase II atrazine metabolism insome fish, insects, mammals and one plant species. Comp Bio-chem Physiol 106:141–149.

Fent K. 1996. Ecotoxicology of organotin compounds. CRC CritRev Toxicol 26.1:1–117.

Grillitsch BBA, Vogl C. 1996. Zur Eignung eines verhaltenstox-ikologischen Prufverfahrens als Ersatzmethode fur Letaltests anFischen in Standardprufverfahren auf Okotoxizitat von Um-weltchemikalien. Forschungsbericht: Institut fur Versuchst-ierkunde Veterinarmedizinische Universitat Wien, A-1210Wien, Josef Baumann-Gasse 1; im Auftrag des Bundesministe-riums fur Wissenschaft und Forschung GZ 49.819/1-Pr/4/95.

Grinwis GCM, Boonstra A, Vandenbrandhof EJ, Dormans JAMA,Engelsma M, Kuiper RV, Vanloveren H, Wester PW, Vaal MA,Vethaak AD, Vos JG. 1998. Short term toxicity of bis (tri-n-butyltin)oxide in flounder (Platichthys flesus)—pathology andimmune funktion. Aquat Toxicol 42:15–36.

Habig W, Pabst MJ, Jacoby WB. 1974. Glutathione-S-transferase:The first step in mercapturic acid formation. J Biol Chem249:1730–1739.

Hartl MGJ, Hutchinson S, Hawkins L. 2001. Organotin and os-moregulation: quantifying the effects of environmental concen-trations of sediment-associated TBT and TPhT on the freshwa-ter-adapted European flounder, Platichthys flesus L. J Exp MarBiol Ecol 256:267–278.

Huuskonen S, Lindstrom-Seppa P, Koponen K, Sashwati R. 1996.Effects of non-ortho-substituted polychlorinated biphenyls (con-geners 77 and 126) on cytochrome p4501a and conjugationactivities in rainbow trout (Oncorhynchus mykiss). Comp Bio-chem Physiol C Toxicol Pharmacol 113:205–213.

Jorgensen EH, Bye BE, Jobling M. 1999. Influence of nutrialstatus on biomarker responses to PCB in the Arctic charr(Salvelinus alpinus). Aquat Toxicol 44:233–244.

Kamjunke N, Schmidt K, Pflugmacher S, Mehner T. 2002. Con-

sumption of cyanobacteria by roach (Rutilus rutilus): useful orharmful to the fish? Freshwater Biol 47:243–250.

Kannan K, Maruya KA, Tanabe S. 1997. Distribution and charac-terization of polychlorinated biphenyl congeners in soil andsediments from a superfund site contaminated with Aroclor1268. Environ Sci Technol 31:1483–1488.

Koponen K, Lindstrom-Seppa P, Kukkonen JVK. 2000. Accumu-lation pattern and biotransformation enzyme induction in rain-bow trout embryos exposed to sublethal aqueous concentrationsof 3,3�,4,4�-tetrachlorobiphenyl. Chemosphere 40:245–253.

Livingstone DR, Nasci C, Sole M, Da Ros L, O’Hara SCM, PetersLD, Fossato V, Wooton AN, Goldfarb PS. 1997. Apparentinduction of a cytochrome P450 with immunochemical similar-ities to CYP1A in digestive gland of the common mussel (Myti-lus galloprovincialis L.) with exposure to 2,2�,3,4�,5�-hexachlo-robiphenyl and Aroclor 1254. Aquat Toxicol 38:205–224.

Lundstrom J, Carney B, Amcoff P, Pettersson A, Borjeson H,Forlin L, Norrgren L. 1999. Antioxidative systems, detoxifyingenzymes and thiamine levels in Baltic salmon that develop M74.AMBIO 28:24–29.

Maguire RJ, Tkacz RJ, Chau YK, Bengert GA, Wong PTS. 1986.Occurrence of organotin compounds in water and sediments inCanada. Chemosphere 15:253–274.

Morcillo Y, Porte C, 1997. Interaction of tributyl- and triphenyltinwith the microsomao monooxygenase system of molluscs andfish from western Mediterranean. Aquat Toxicol 38:35–46.

Oberdorster E, Rittschof D, McClellan-Green P. 1998. Inductionof cytochrome P450 3A and heat shock protein by tributyltin inblue crab, Callinectus sapidus. Aquat Toxicol 41:83–100.

Otto DME, Moon TW. 1996. Phase I and II enzymes and antiox-idant responses in different tissues of brown bullheads fromrelatively polluted and non-polluted systems. Arch EnvironContam Toxicol 31:141–147.

Pang S, Cao JQ, Katz BH, Hayes CL, Sutter TR, Spink DC. 1999.Inductive and inhibitory effects of non-ortho-substituted poly-chlorinated biphenyls on estrogen metabolism and human cyto-chromes P450 1A1 and 1B1. Biochem Pharmacol 58:29–38.

Perez Lopez M, Novoa-Valinas MC, Melgar-Riol MJ. 2002. Glu-tathione S-transferases cytosolic isoforms as biomarkers ofpolychlorinated biphenyls (Aroclor 1254) experimental contam-ination in rainbow trout. Toxicol Lett 136:97–106.

Pflugmacher S, Steinberg C. 1997. Activity of phase I and phase IIdetoxication enzymes in aquatic macrophytes. Angew Bot 71:144–146.

Pflugmacher S, Ame V, Wiegand C, Steinberg C. 2001. Cyanobac-terial toxins and endotoxins—their origin and their ecophysi-ological effects in aquatic organisms. Wasser Boden 53:15–20.

Rice JA. 1990. Bioenergetics modelling approaches to evaluationof stress in fishes. American Fisheries Society Symposium,Bethesda, MD 8:80–92.

Safe S. 1990. Polychlorinated biphenyls (PCBs), dibenzo 1,4 di-oxins (PCDD), dibenzofurans (PCDFs) and related compounds:Environmental and mechanistic considerations which supportthe development of toxic equivalent factors (TEFs). Crit RevToxicol 21:51–88.

Schulte-Oehlmann U, Stroben E, Fiorini P, Oehlmann J. 1996.Beeintrachtigung der Reproduktionsfahigkeit limnischer

XENOBIOTICS INFLUENCE CARP SWIMMING/GST 469

Vorderkiemerschnecken durch das Biozid Tributylzinn (TBT).In: Lozan JL, Kausch H, editors. Warnsignale aus Flussen undAstuaren: wissenschaftliche Fakten. Berlin: Parey. p 249–255.

Siegel S. 1985. Nichtparametrische statistische Methoden. Fach-buchhandlung fur Psychologie. Eschborn bei Frankfurt amMain.

Spieser H. 2001. Quantitative behavior analysis—a new approachto the challenges of environmental toxicology. In: ButterworthFM, Gunatilaka A, Gonsebatt ME, editors. Biomomitors andbiomarkers as indicators of environmental change: 2. A hand-book. New York: Kluwer Academic/Plenum Publishers.

Staaks G. 1996. Experimental studies on temperature preferencebehaviour of juvenile Cyprinids. Limnologica 26:165–177.

Steinberg CEW, Lorenz R, Spieser OH. 1995. Effects of atrazineon swimming behavior of zebrafish, Brachydanio rerio. WaterRes 29:981–985.

Svobodova Z, Vykusova B, Machova J, Hrbkova M, Groch L.1994. Long term effects of PCBs on fish. In: Muller R, Lloyd R,editors. Sublethal and chronic effects of pollutants on freshwaterfish. Oxford, UK: Fishing News Books.

Warthold R. 1980. Untersuchungen zur Beeinflussung derSchwimmaktivitat und des systeminternen Zustandes vonKarpfengruppen (Cyprinus carpio L.) durch das Zeitmuster derFutterung, das Lichtregime und die Besatzdichte in Korrelationzur Zuwachsrate [Diss. A]. Humboldt-Universitat, Berlin.

Whyte JJ, Jung RE, Schmitt CJ, Tillitt DE. 2000. Ethoxyresorufin-O-deethylase (EROD) activity in fish as a biomarker of chem-ical exposure. Crit Rev Toxicol 30:347–570.

Wiegand C, Pflugmacher S, Giese M, Frank H, Steinberg CEW.2000. Uptake, toxicity and effects on detoxication enzymes ofatrazine and trifluoroacetate in embryos of zebrafish. EcotoxicolEnviron Saf 45:122–131.

470 SCHMIDT ET AL.