Aquatic Toxicology 73 (2005) 79–90

Sex and low-level sampling stress modify the impacts of sewageeffluent on the rainbow trout (Oncorhynchus mykiss)

immune system

Birgit Hoegera, Bettina Hitzfeldb, Bernd Kollnerc, Daniel R. Dietricha,Michael R. van den Heuveld,∗

a Environmental Toxicology, University of Konstanz, Germanyb Swiss Agency for the Environment, Forests and Landscape, Berne, Switzerland

c Friedrich-Loeffler-Institute, Greifswald-Insel Riems, Germanyd Forest Research, Private Bag 3020, Sala Street, Rotorua, New Zealand

Received 10 January 2005; received in revised form 21 March 2005; accepted 22 March 2005

Abstract

The objective of the present study was to investigate the influence of chronic exposure to municipal sewage treatment efflu-ent at environmentally relevant concentrations on immune parameters in rainbow trout (Oncorhynchus mykiss), including theassessment of potential differences in reactivity between sexually mature male and female fish. Trout were exposed to 1.5 and15% (v/v) secondary treated municipal sewage effluent for 32 weeks. Fish were injected intra-peritoneally either with inactivated

pling.s inaledays waseripheralrophagesd day offluent. Ins, can leadthat bothle fish andto detect.

e

Aeromonas salmonicidato simulate an infection or with PBS as control for this immune challenge 6 weeks prior to samExposure to effluent resulted in a decrease inA. salmonicida-specific serum antibody level and blood lymphocyte numbermature females, but not in male fish. Injection ofA. salmonicidaresulted in enhanced serum lysozyme activity in mature mtrout, which were not exposed to effluent. This stimulating effect ofA. salmonicidacould not be found in effluent-exposetrout, again potentially revealing a suppressive effect of the effluent. An influence of sampling fish on two consecutive dobserved in many immune parameters, most likely reflecting handling stress. Leucocyte and lymphocyte numbers in pblood were consistently lower in male and female fish on the second sampling day. Phagocytosis in head kidney macfrom male trout was also influenced by sampling day, whereby a stimulation of this reaction occurred on the seconsampling. Liver mixed function oxygenase activity was found to be enhanced in mature male trout exposed to 15% efconclusion, the study showed, that exposure to sewage treatment plant effluent, in surface water relevant concentrationto potentially adverse effects on selected immune reactions in rainbow trout. However, this study also demonstratedhandling stress and the sex of mature fish have distinct influences on the immune response detected in male and femaare likely to influence measured immune parameters to the extent that subtle effluent induced changes may be difficult© 2005 Elsevier B.V. All rights reserved.

Keywords:Rainbow trout;Oncorhynchus mykiss; Fish immune system; Immunotoxicology;Aeromonas salmonicida; Handling stress; Sewagtreatment plant; Effluent

∗ Corresponding author. Tel.: +64 7 343 5899; fax: +64 7 343 5695.E-mail address:mike.vandenheuvel@forestresearch.co.nz (M.R. van den Heuvel).

0166-445X/$ – see front matter © 2005 Elsevier B.V. All rights reserved.doi:10.1016/j.aquatox.2005.03.004

80 B. Hoeger et al. / Aquatic Toxicology 73 (2005) 79–90

1. Introduction

In the last two decades sewage treatment efflu-ents have been found to contain a broad array ofanthropogenic substances, including pharmaceuticalresidues (Daughton, 2001; Kolpin et al., 2002), chem-icals contained in household products, such as vari-ous derivatives of alkylphenol (Bennie, 1999), as wellas natural and synthetic hormones (Desbrow et al.,1998). In spite of usually being present in low con-centrations, some of these environmental pollutantshave the potential to affect the health of aquatic or-ganisms, due to their specificity for eliciting effectson certain biochemical functions. Pharmaceuticals inparticular are designed to specifically influence sub-tle biochemical mechanisms and might thus affectsimilar functions in non-target aquatic species in thelow concentrations found in aquatic environments.Hormone-like activity of several low-level pollutantsin our surface waters has already been investigated(Ashfield et al., 1998), while effects on immune mech-anisms have not received a similar level of attention.However, as some of the commonly found pollutantsare known to alter immune reactions or, in the caseof some pharmaceuticals, are even designed to influ-ence the immune system, it should be assumed thatnot only reproductive-endocrine mechanisms can beaffected in aquatic organisms, but that immune reac-tions might also be a target for some of the aquaticlow-level pollutants. Effects on various immune pa-r verals e ef-fla d eta l-lB retp ef-fl as T-a ymea . Inas asei rs inb1

A key element to assessing the competency of theimmune system is to examine responses with and with-out the presence of an immune activator. One suchapproach is to expose the organism to inactivatedpathogens. The infection model applying inactivatedA.salmonicidain trout is known to result in an activationof different leukocyte populations e.g. monocytes andB-cells leading to a specific antibody response againstbacterial antigens (Kollner and Kotterba, 2002). Fur-thermore, MHC II expression was found to be increasedin monocytes, B-lymphocytes and a part of the throm-bocytes, indications of a general stimulation of the en-tire immune system (Stachowski et al., 2004).

Stress is known to be an important factor in fishhealth and is also a potential confounding factor in ex-periments investigating the immune functions of fishes.Effects of stress on immune parameters in fish havebeen investigated in various studies and e.g. a sup-pressed antibody response against sheep red blood cellsin Atlantic salmon (Salmo salar) (Einarsdottir et al.,2000) and reduced resistance againstTrypanoplasmaborrelii infection in carp (C. carpio) (Saeij et al., 2003)has been detected after rearing stress due to repeatedwater level reduction and daily handling, respectively.Acute stress, however, has also been found to en-hance specific immune reactions, as has been shownfor lysozyme activity in rainbow trout (O.mykiss) aftera 10 min handling stress (Demers and Bayne, 1997).Depending on the specific situation, handling stressmay thus enhance or oppose possible adverse effectso

ex-p atert owt ep-r bilityo wl-e con-f ake,l outi d ofe ren-t im-m andh ando uc-t as of

ameters in fish have already been shown for seubstances, well known to be present in sewaguent, including PCBs (Khan, 2003), PAHs (Weeksnd Warinner, 1986; Karrow et al., 2001; Reynaul., 2003), pesticides (Dunier and Siwicki, 1993; Ga

oway and Handy, 2003) and antibiotics (Lunden andylund, 2000). In the only study examining exposu

o wastewater in the field, caged carp (Cyprinus car-io) were exposed to river water receiving sewageuent (Price et al., 1997). Exposed fish displayedignificant reduction in the proliferative ability ofnd B-lymphocytes, and a decrease in serum lysozctivity in comparison to fish from a reference sitelaboratory experiment, exposure of goldfish (Carra-ius auratus) to treated sewage resulted in a decren erythrocyte, granulocyte and lymphocyte numbelood and lower blood cell phagocytic activity (Kakuta,997).

f water pollutants on immune reactions in fish.The aim of the present study was to determine if

osure to a sewage effluent from a modern wastewreatment plant affects immune functions in rainbrout (O. mykiss). This species was chosen as a resentative of teleosts because of the large availaf validated test methods and immunological knodge. In order to provide relevance and eliminate

ounding environmental factors such as energy intong-term exposure of sexually maturing rainbow trn a mesocosm facility was chosen as the methoxposure. General immune parameters, like diffeial blood cell counts, reactions of the unspecificune system, including serum lysozyme activityead kidney macrophage activity (phagocytosisxidative burst), as well as specific antibody prod

ion against the trout pathogenAeromonas salmonicidalmonicidawere determined. The high relevance

B. Hoeger et al. / Aquatic Toxicology 73 (2005) 79–90 81

immune parameters as endpoints chosen in this chronicexposure study with sewage effluent is based on thepotential of imminent immune suppression to lead tohigher disease susceptibility, thus likely endangeringthe health of fish populations.

2. Material and methods

2.1. Experimental design

Two-year-old rainbow trout (Oncorhynchusmykiss)were exposed for 32 weeks to a nominal concentrationof 1.5% (v/v) or 15% (v/v) municipal sewage treat-ment plant (STP) effluent in 12,000 L tanks containing50 individuals. Final treated effluent was obtained froma municipal sewage treatment plant located in Rotorua,New Zealand every 2 days. This STP employs a pre-treatment step with stop screens and a grit trap, a pri-mary treatment step with sedimentation and secondaryactivated sludge treatment (Bardenpho Reactor).

Activated carbon de-chlorinated tap water (aquifersource) was used as the diluent and as the referencetreatment. Two replicate tanks were used for eachtreatment. Trout exposure tanks were provided witha constant water flow of 10 L/min, which resulted ina 95% replacement time of approximately 45 h. Dis-solved oxygen, pH and conductivity (Radiometer Pa-cific, Auckland, New Zealand) in the fish exposuretanks and in undiluted effluent were measured daily.A ld-i ningd du-r ba-s et -flec on-d cu-l nkso ationw tt -e 7 and1t ely.A ar

type tags (HallPrint Pty Ltd., Holden Hill, SA, Aus-tralia). Weight and length for each tagged trout weremeasured at the start of the experiment. Trout werefed daily with commercial feed pellets (Reliance StockFood, Dunedin, New Zealand) at a ration of 0.7% of thewet body weight and this ration was adjusted monthlyto compensate for growth. The exposure was started onSeptember 22, 2001, with immature fish and continuedfor 32 weeks. After an exposure period of 26 weekstrout were anesthetized with ethyl-3-aminobenzoatemethanosulfonate (MS222; Fluka, Switzerland) and1 mL of blood was taken by syringe from the caudalvein. Fish were then either injected with formalde-hyde inactivatedA. salmonicida salmonicidastrainMT 423 antigen (1× 108 cells in 25�L per 100 gbody weight) (kindly provided by Gunter Kotterba,Friedrich-Loeffler-Institute, Insel Riems, Germany), orwith a corresponding volume of phosphate balancedsalt solution (PBS) as a control for the injection. Anti-gen preparation ofA. salmonicidafollowed the descrip-tion byKollner and Kotterba (2002). Fish were exposedfor a further 6 weeks and were sampled between May 6and 14, 2002, just prior to the natural spawning periodfor this strain of trout. Female trout were sampled overtwo consecutive days on May 6 and 7. Male trout weresampled over two consecutive days on May 13 and 14,one week later than the females in order to allow themto recover from the short-term stress due to removal offemales from the tanks. At the time of sampling, nor-mal aggressive feeding behaviour had resumed in them s ap arso uner ow-e t the2 on-j pleda

2

oodw mearw plesw sam-p ed at1 st ey

dditional aeration was provided in the effluent hong tanks and the trout exposure ponds maintaiissolved oxygen above 90% saturation for theation of the experiment as measured on a dailyis. The average pH-values (±S.D.) in the exposuranks were 7.21± 0.25 and 7.22± 0.25 in the 15% efuent tanks, 7.15± 0.28 and 7.20± 0.28 in the 1.5%ffluent tanks and 7.13± 0.29 and 7.13± 0.28 in theontrol tanks. Conductivity in each tank, diluent cuctivity, and effluent conductivity was used to cal

ate the actual effluent concentration in the fish tan a daily basis. The actual mean effluent concentras 13.48± 3.25 and 12.91± 3.51 in the 15% effluen

anks and 1.45± 0.59 and 1.59± 0.64 in the 1.5% efflunt tanks. Water temperature ranged between 12.6.7◦C, 12.7 and 16.9◦C and 12.9 and 17.5◦C in con-

rol water, 1.5% effluent and 15% effluent, respectivll trout were tagged with individually numbered T-b

ale trout. This particular strain of trout typically haroportion of individuals that do not mature at 3 yef age. Immature trout were not examined for immesponses at the termination of the experiment. Hver, since blood was removed from these fish a6 weeks time-point, lysozyme was examined in c

unction with the maturing males and females samt this time-point.

.2. Sampling

Fish were sacrificed with a blow on the head. Blas removed by caudal puncture and a blood sas prepared immediately. Blood for serum samas collected in untreated vacutainers. Theseles were kept on ice and subsequently centrifug000×g (10 min, 4◦C) to obtain serum, which wa

hen kept at−80◦C pending analysis. Head kidn

82 B. Hoeger et al. / Aquatic Toxicology 73 (2005) 79–90

for macrophage preparation was collected and kepton ice in Leibovitz’s L-15-medium (Invitrogen, Auck-land, New Zealand), containing 53 mU heparin sodiumsalt/mL (Sigma, St. Louis, USA) and 100 U/mL ofPenicillin/Streptomycin (P/S) (Invitrogen, Auckland,New Zealand). A 1 g sample of liver was frozen inliquid nitrogen for subsequent determination of 7-ethoxyresorufin-O-deethylase (EROD) activity.

2.3. Immunological parameters

Blood smears were stained according to Pappen-heim (May-Gruenwald/Giemsa staining). Pictures ofthe blood smears were taken with an SV Micro SoundVision color camera (Sound Vision Inc., Boston, USA)on a microscope (Zeiss Axiolab) using Axio Vi-sion Version 2.0.5. (Carl Zeiss Vision GmbH, Hall-bergmoos, Germany). A total of 1500 cells per slidewere counted. The different blood cell populationswere expressed in percent of total cells counted.The preparation of head kidney macrophages wasperformed according toSecombes (1990)using aPercoll gradient (Sigma, St. Louis, USA). The re-sulting cell suspension was adjusted to a density of1× 106 cells/mL in L-15 medium, containing P/S,and 100�L/well of this suspension was seeded into96-well black fluorometer plates (BMG Labtechnolo-gies, Offenburg, Germany). After incubation at 18◦Cfor 90 min to allow attachment of macrophages, me-dia was removed by inverting the plate on a papert

o -s SA)w lls,w tiona thea et cu-b platefl ies,O is-s

/mLp uis,U ntot ac-t 2

7′-dichlorodihydrofluorescein diacetate (H2DCFDA)(Molecular Probes, Eugene, USA) at a concentrationof 10�g/mL. The time course of H2O2-productionwas measured in a microplate fluorometer (485 nm ex-citation, 520 nm emission). The slope was obtainedfrom data measured from 4 to 10 min. Results werecalculated as pmol 2′-7′-dichlorofluorescein (DCF)(Acros, Schwerte, Germany) produced per minute perwell.

The method for measuring serum lysozyme fol-lows the description ofEllis (1990) (turbimetric as-say). Briefly, 950�L M. lysodeikticussolution (Sigma,Steinheim, Germany) (0.5 mg/mL in 0.05 M sodiumphosphate buffer, pH 6.2) were pipetted in a cuvetteand measured in a spectrophotometer at 530 nm, fol-lowed by the addition of 50�L of serum or plasma.Absorption was then measured after 30 s, 1, 2, 3, 4and 4.5 min and lysozyme activity was expressed asdecrease of optical density (OD) per min.

The ELISA for the detection ofA. salmonicida-specific antibodies in trout serum followed the methoddescribed byKollner and Kotterba (2002). Briefly,Maxisorp ELISA plates (Nunc, Roskilde, Denmark)were coated with 20�g/mL A. salmonicida salmoni-cidastrain MT 423 antigen (kindly provided by GunterKotterba, Friedrich-Loeffler-Institute, Insel Riems,Germany) overnight at 4◦C. Plates were blocked with1% ovalbumin (Sigma, Steinheim, Germany) in PBSfor 1 h at 20◦C, followed by washing three timeswith PBS containing 0.05% Tween 20 (T-PBS) (CarlR ionw for1 in-c antit -i ate(t car-r c-c uis,U itiono nmi nts,G forI ablec erem asr

owel.For the phagocytosis assay, a volume of 100�L

f a 250�g/mL fluorescein-labeledE. coli suspenion (K-12 strain, Molecular Probes, Eugene, Uas added to each well including eight blank wehich did not contain macrophages. After incubat 18◦C for 2 h, media was removed, followed byddition of 100�L trypan blue solution (0.025%). Th

rypan blue solution was removed after 1 min of ination and fluorescence was measured in a microuorometer (Polarstar Galaxy, BMG Labtechnologffenburg, Germany; 485 nm excitation, 520 nm emion).

For the oxidative burst assay, either 200 nghorbol-myristate-acetate (PMA) (Sigma, St. LoSA) in HBSS or HBSS alone was pipetted o

he cells. Measurement of the oxidative burst reion was started 5 min later by the addition of′-

oth GmbH, Karlsruhe, Germany). After incubatith diluted serum samples (1:4000 in T-PBS)h at 20◦C and another wash cycle, plates wereubated with monoclonal antibody 4C10 (mouse-rout immunoglobulin M) for 1 h at 20◦C. After washng, goat-anti-mouse IgG/IgM-peroxidase conjugPierce, Rockford, USA) was added for 1 h at 20◦C. Af-er washing three times with T-PBS, detection wasied out with 3,3′,5,5′-tetramethylbenzidine (TMB) aording to manufacturers description (Sigma, St. LoSA). The colour reaction was stopped by the addf 1 M H2SO4 and absorption was measured at 450

n an SLT plate reader 340 ATTC (SLT Labinstrumeroedig, Austria). As no standards were available

gM determination, results are given as OD. To enomparison without a standard curve all samples weasured in parallel in a single ELISA run, which w

epeated twice.

B. Hoeger et al. / Aquatic Toxicology 73 (2005) 79–90 83

2.4. Liver 7-ethoxyresorufin-O-deethylase(EROD) activity

Hepatic mixed function oxygenase (MFO) activ-ity was assessed in liver post-mitochondrial super-natant (PMS) as EROD activity using a fluorescenceplate-reader method. Liver extracts were homogenisedin 0.1 M phosphate buffer containing 1 mM EDTA,1 mM dithiothreitol, and 20% glycerol at pH 7.4. Ho-mogenates were centrifuged at 9000×g to obtainPMS. EROD reaction mixture contained 0.1 M HEPESbuffer pH 7.8 (Sigma, St. Louis, MO, USA), 5.0 mMMg2+, 0.5 mM NADPH (Applichem, Darmstadt, Ger-many), 1.5�M 7-ethoxyresorufin (Sigma, St. Louis,MO, USA), and about 0.5 mg/mL of PMS protein. Re-sorufin was determined using 544 nm excitation and590 nm emission filters. Protein content was estimatedfrom fluorescamine (Sigma, St. Louis, USA) fluo-rescence (390 nm excitation, 460 nm emission filters)against bovine serum albumin (Sigma, St. Louis, USA).

2.5. Statistics

Immune parameters in mature male and femaletrout were analysed independently using ANOVA withdata appropriately transformed to meet the assump-tions of homogeneity of variance and normality. Thevariability of replicate tanks was tested as a vari-able in all ANOVAs and did not add significant vari-ability for any of the endpoints. Sampling day,A.s nyp s ina abil-i ari-a testf ther VAm d too vari-a entA y oft us-i ly-s plec us-iT ticald

3. Results

External examination of trout at 26 and 32 weeksindicated no obvious signs of external opportunisticinfections such as fin erosion, fungus, or other skin le-sions, though fins did occasionally show damage con-sistent with confinement due to aggression or abrasionon the tanks. Slightly increased mortality was experi-enced in the 15% effluent during the summer months,whereby 5% of the initially 100 trout of this groupdied.

3.1. Innate immune measurements

Data on the different immune parameters investi-gated were analysed for effects of effluent exposureandA. salmonicida-injection as well as consequencesof sampling on two consecutive days possibly lead-ing to handling stress in remaining fish after the firstsampling day. Differences in reactivity of immunefunctions between male and female trout were alsoassessed.

Leucocyte and lymphocyte numbers in peripheralblood were not influenced byA. salmonicida-injectionin either males or females. Leucocyte numbers as wellas lymphocyte numbers were, however, consistentlyand statistically significantly lower in male and femalefish on the second sampling day compared to day 1(Fig. 1). There were consistent trends for reductionsin lymphocytes and total leucocytes with increasinge nif-i oodc tesi am-p ont tesa xpo-s tan

alessa pat-t 2 asi flu-e roms ent-e ox-i ant

almonicida-injection, effluent concentration and aossible interactions were examined as variablell subsequent analyses. Where no significant vari

ty due to the tested variables was found, that vble was eliminated from the model. Dunnett’s

or post hoc comparisons of treatment groups toeference group was performed on the final ANOodel. Where statistical interactions were observeccur, the data was split based on the interactingble (sampling day or injection) and an independNOVA was performed for each possible categor

hat particular variable. EROD data were comparedng a non-parametric Kruskal–Wallis one-way anais of variance with Bonferroni adjustment for multiomparisons. All statistical testing was completedng the SYSTAT® software package (Wilkinson, 1990).he level of confidence used in determining statisifferences for all analyses wasα = 0.05.

ffluent concentration, however, a statistically sigcant effect of effluent exposure on peripheral blell numbers could only be found for lymphocyn females on the first but not on the second sling day. Also, no significant effects were found

he proportion of circulating thrombocytes, monocynd granulocytes in either sex due to effluent eure,A. salmonicida-injection, or sampling day (daot shown).

For the oxidative burst assay, males and femhowed very different patterns of response (Fig. 2and b). Females showed an entirely different

ern of response on day 1 as opposed to dayndicated by a significant interaction between efnt exposure and sampling day. When data fampling days were examined separately, effluxposed females showed significantly increased

dative burst on day 1. There was a non-signific

84 B. Hoeger et al. / Aquatic Toxicology 73 (2005) 79–90

Fig. 1. Peripheral blood leucocyte numbers in female (a) and male trout (b) and lymphocyte numbers in female (c) and male fish (d) after 32weeks of exposure to effluent. Female and male fish were each sampled on two consecutive days (days 1 and 2) and data was accordinglyanalysed for effects of effluent exposure and sampling day. Shown are mean values with standard error of the means (S.E.M). Data was testedusing a three-way ANOVA with Dunnett’s post-test.n≥ 6; *p< 0.05.

trend for oxidative burst to decrease with increasingeffluent dose in females on day 2. Oxidative burstin males was not altered by effluent exposure, butreactivity on day 2 was significantly lower than onday 1.

As with oxidative burst, kidney macrophage phago-cytosis showed very different patterns of response tosampling day in males and females (Fig. 2c and d).Phagocytosis in females was not affected by samplingday, whereas in males, phagocytosis greatly increasedon day 2 of sampling. There were no effects of sewageeffluent on phagocytosis in either males or females.

Serum lysozyme activity was measured after 26weeks of exposure and at the end of the experimentafter 32 weeks. Immature fish (including male and fe-male fish) displayed a significant decrease in serumlysozyme activity after exposure to 1.5% or 15% ef-fluent for 26 weeks (Fig. 3a). In mature female troutat 26 weeks, exposure to 1.5 and 15% effluent also re-sulted in decreased lysozyme activity, however, only

statistically significant in the group exposed to 1.5%effluent. In mature male fish, exposure to effluenthad no effect on serum lysozyme activity after 26weeks of exposure. Upon termination of the experi-ment (32 weeks) only data for mature male trout wasanalysed, as immature fish were not sampled at thistime-point, and the majority of serum samples fromfemale fish caused precipitation in theM. lysodeik-ticus solution used for the lysozyme assay. Expo-sure to 1.5% or 15% STP effluent for 32 weeks didnot have an effect on serum lysozyme activity. How-ever, in control fish exposed to de-chlorinated tap wa-ter, i.p.-injection with the inactivated trout pathogenA. salmonicidaresulted in a significant increase inserum lysozyme activity compared to PBS-injected fish(Fig. 3b). In the fish exposed to effluent, no differencein lysozyme activity between pathogen-injected andsham-injected trout could be found. Serum lysozymeactivity was not observed to be sensitive to handlingstress.

B. Hoeger et al. / Aquatic Toxicology 73 (2005) 79–90 85

Fig. 2. Head kidney macrophage oxidative burst in female (a) and male trout (b) and macrophage phagocytosis in female (c) and male fish (d)after 32 weeks of exposure to effluent. Female and male fish were each sampled on two consecutive days (days 1 and 2) and data was accordinglyanalysed for effects of effluent exposure and sampling day. Shown are mean values with S.E.M. Data was tested using a three-way ANOVA withDunnett’s post-test,n≥ 8 for (a) and (c);n≥ 5 for (b) and (d);*p< 0.05;*** p< 0.001.

3.2. Aeromonas salmonicida-specific antibodyELISA

Trout sham-injected with PBS and kept together inthe same exposure tanks withA. salmonicida-injectedfish did not show detectable serum levels of anti-A.salmonicidaantibodies, excluding a possible cross-contamination. As no detectable antibody productionwas evident, samples from sham-injected fish were ex-cluded from further analysis. Male and female troutagain showed a different pattern of response. In maturefemale trout, exposure to 1.5% or 15% STP effluent re-sulted in lower serum levels of specific antibodies, pro-duced against inactivatedA. salmonicida.The decreasein serum antibody levels was, however, only statisti-cally significant in the group exposed to 1.5% effluent(Fig. 4). Mature male trout did not display significantlyaltered levels of specific antibodies after exposure to ef-fluent compared to tap water exposed fish and unlikeantibody levels in females this measure tended to in-

crease with effluent concentration. Serum level of anti-A. salmonicidaIgM were not observed to be influencedby stress due to consecutive sampling days.

3.3. Liver EROD activity

Exposure to 15% effluent over 32 weeks resulted ina significant increase in liver EROD activity in maturemale fish compared to control fish (Fig. 5). In maturefemale fish, no effect of STP effluent on liver ERODactivity could be shown.

4. Discussion

Chronic exposure of rainbow trout to treated sewageeffluent resulted in slightly increased mortality, pre-sumably due to increased ammonia levels occasionallydetected in the effluent in the course of regular moni-toring at the sewage treatment plant (Tim Charleson,

86 B. Hoeger et al. / Aquatic Toxicology 73 (2005) 79–90

Fig. 3. Lysozyme activity in serum samples, taken after exposure toeffluent for: (a) 26 weeks (prior to injection withA. salmonicida,n≥ 21) and (b) after 32 weeks (n≥ 7), where data was analysed foreffects of effluent exposure andA. salmonicida-injection.Shown aremean values with S.E.M. Data was tested using one-way ANOVAwith Dunnett’s post-test;*p< 0.05;** p< 0.01.

Fig. 4. A. salmonicida-specific antibody level in trout serum sam-ples after exposure to effluent for 32 weeks. Data was tested usingone-way ANOVA with Dunnett’s post-test. Shown are mean valueswith S.E.M. Male trout:n= 12 for the groups exposed to 0 and 15%effluent;n= 7 for the 1.5% effluent group. Female fish:n= 12 for thegroups exposed to 0 and 1.5% effluent;n= 8 for the 15% effluentgroup;*p< 0.05.

Fig. 5. Liver EROD activity in trout after exposure to STP ef-fluent for 32 weeks. Shown are median values with S.E.M. Datawas tested with Kruskall–Wallis non-parametric one-way ANOVA.n≥ 17; *** p< 0.001.

Rotorua District Council, personal communication).Exposure of rainbow trout to STP effluents induced asuppression of peripheral blood lymphocyte numbers,serum lysozyme activity and specific antibody produc-tion, while the production of reactive oxygen specieswas enhanced only in effluent-exposed female fish.

Significantly reduced peripheral blood lymphocytenumbers were also observed in an earlier experimentwith juvenile rainbow trout exposed to high concen-trations of effluent from the same STP for 28 days(Hoeger et al., 2004b). Similarly, Kakuta (1997)re-ported lower lymphocyte numbers in goldfish (C. au-ratus) after exposure to treated effluent in a laboratoryexperiment. A possible cause for the decrease of lym-phocyte numbers could be polyaromatic hydrocarbons(PAH), as has been described before byKhan (2003),who found lower circulating lymphocyte numbers inwinter flounder (Pleuronectes americanus) sampled ata PAH-polluted boat wharf. The presence of PAHs inthe effluent used for the study at hand has been sug-gested by analysis of bile samples of trout exposed tohigher concentrations of effluent from the same STPfor 28 days (Hoeger et al., 2004b). As PAHs are knownto be strong activators of cytochrome P450 1A1, thesignificant induction of liver EROD activity in effluent-exposed trout, whether as juveniles for 28 days (Hoegeret al., 2004b) or as adult male fish for 32 weeks (thisstudy), further supports the assumption that these fishwere exposed to PAH via the sewage treatment efflu-ent used. At first sight it seems contradictory to explaint male

he reduced lymphocyte numbers observed in fe

B. Hoeger et al. / Aquatic Toxicology 73 (2005) 79–90 87

trout by exposure to PAHs while an increase in ERODactivity was not demonstrated in mature females in thepresent study. However, in this case the lack of ERODinduction does not exclude exposure to PAHs, as ERODactivity has previously been found to be suppressed inreproductively mature female trout, presumably due toan estrogen dependent mechanism (van den Heuvel andEllis, 2002).

The decreased ability of mature female trout toproduceA. salmonicida-specific antibodies was mostlikely a direct consequence of the lower circulatinglymphocyte numbers in peripheral blood, after ex-posure to STP effluent. Immunohistology in spleensamples from the same exposure experiment revealedan activation of B-cells in effluent-exposed femaletrout, implied by higher surface immunoglobulin Mexpression on this cell type (Hoeger et al., 2004a). Atfirst sight, this observation seems to be in contradic-tion to the observed lower specific serum antibody lev-els. However, the activation of B-lymphocytes in thespleen might be the result of effluent-induced unspe-cific stimulation resulting in a polyclonal stimulationof B-cells and, hence, appear unrelated to the produc-tion of A. salmonicida-specific antibodies. A possi-ble effluent-induced stimulatory effect on head kidneymacrophages was observed, as suggested by the in-creased oxidative burst in female trout. Whether theobserved stimulation of B-lymphocytes in spleen isa direct consequence of antigen presentation by thesemacrophages following effluent exposure cannot be de-t dy.

t-e ymea hers ef-fl oft In-d lg activ-i hef eo sivee asa , i.e.a tedi ymea ichw ,

1997). PAHs have moreover been considered to bethe active component of creosote leading to reducedplasma lysozyme level in rainbow trout (Karrow et al.,1999). Whether the suppression of lysozyme activitywas due to the suggested concomitant PAH exposurewith the effluents cannot be determined, as chemicalanalysis of PAHs in the effluent used were not con-ducted. However, the assumed effect of PAHs in theeffluent on lysozyme activity is consistent with theother seemingly PAH-related immune modulations de-scribed above and thus suggests that PAHs could be theimmune-suppressive component in the STP effluentstested in this study.

The seeming lack of effect ofA. salmonicidaon im-mune parameters other than lysozyme activity mightbe due to the time-point chosen for sampling (6 weeksafter injection). It is possible that fast reacting in-nate immune functions had actually been stimulated bypathogen-injection, but have already returned to base-line values at the time of sampling. Effects expectedby A. salmonicidamay thus have simply been misseddue to a time schedule inappropriate to examine fastreacting immune parameters.

Effects of different kinds of stress on immune pa-rameters have been studied extensively, especially incorrelation with the role of cortisol as an importantstress mediator (Narnaware et al., 1994; Espelid et al.,1996; Rotllant et al., 1997). However, in studies onpollution effects, possible consequences of handlingstress, due to sampling procedures are often ignored. Int ffectso plingd uco-c alea und.I g ind dlings l re-l es inp ouss -t cor-t al.,2

ces-s canb text,o rded

ermined with the data obtained in the present stuA. salmonicida-injected trout in the effluen

xposed groups did not display enhanced lysozctivity compared to the PBS-injected fish, furtuggesting an immune-suppressive effect of thisuent, in this case specifically on the capabilityhe fish to express this bacteriolytic enzyme.eed, A. salmonicida-injected trout in the controroups showed the expected enhanced lysozyme

ty (Moyner et al., 1993), thus demonstrating that tormaldehyde inactivatedA. salmonicidawas capablf stimulating an immune response. The suppresffect of STP effluent on serum lysozyme activity wlso observed at an earlier time-point in this studyt the 26-week time-point in non-pathogen-stimula

mmature and female fish. Suppression of lysozctivity has also been shown previously in carp, where held in an effluent receiving river (Price et al.

he present study, results were also analysed for ef handling stress due to the two consecutive samays. In agreement with the literature, decreased leyte and specifically lymphocyte numbers in both mnd female fish on the second sampling day were fo

t could be speculated that the mechanism resultinecreased lymphocyte numbers is based on hantress on the first sampling day leading to cortisoease and subsequently apoptosis of B-lymphocyteripheral blood. This idea is derived from previtudies in carp (C. carpio) revealing induction of apoposis in peripheral blood lymphocytes caused byisol (Verburg-van Kemenade et al., 1999; Saeij et003).

Barton (2002)has stated that stress is not nearily detrimental to fish, as long as homeostasise regained by adaptive mechanisms. In this connly severe or long-lasting stressors may be rega

88 B. Hoeger et al. / Aquatic Toxicology 73 (2005) 79–90

as detrimental to fish health. However, little is knownabout the combined effects of different kinds of stres-sors, like e.g. handling stress in combination with poorwater quality. Depending on the immune parameter in-vestigated and the stressors that act on the fish, effectsmay be additive or opposed. In the case of additive ef-fects, adverse effects of aquatic pollution on fish healthbecome exceedingly likely. In the field of environmen-tal toxicology, special consideration should be givento the distinction between the actual pollution effectsand the influence of stress, which can hardly be com-pletely avoided in any experimental procedure. Somealleged toxicological effects of a certain environmen-tal contaminant might actually stem from, or at least beinfluenced by, unconsidered stressors.

The present study also underlines the dependenceof different immune parameters in salmonids on sex-ual maturation and subsequently the sex of the inves-tigated fish, as demonstrated by differences in sensi-tivity of serum lysozyme activity to effluent exposurein immature, male and female fish, as well as sensi-tivity of specific antibody production against effluentexposure in female, but not male trout. Differences inthe response to effluent between male and female fishhave also been found for macrophage activity, wherebyonly female fish but not male trout showed increasedproduction of reactive oxygen species. A similar ob-servation has been described byFournier et al. (1998),who found that exposure of mummichogs (F. heterocli-tus) to pulp mill effluent resulted in a decreased oxida-t alefi sen-s udya temo tisol,g hers( essr beeno onsesw han-d -f nsis-t ut tos

c e fe-m fter3 lop-

ment. Substances like lectin, C-reactive protein (CRP)and lysozyme are present in fish eggs as a defenseagainst infection (Yousif et al., 1994a, 1994b; Nuno-mura, 1991) and are likely to be present in higheramounts in blood upon maturation of female fish. Thecapacity to agglutinate bacteria has e.g. been shownfor egg lectin in coho salmon (O. kisutch) (Yousif etal., 1994b).

While, in recent years, the applicability of endocrineparameters for the assessment of aquatic pollution hasbeen widely investigated and discussed, immunotoxi-cology in fish is still undeveloped due to the lack in theunderstanding of basic immunology in fishes and theimmense complexity of the immune system. The ex-perimental design, needed to enable the assessment ofimmune parameters, is complicated by various charac-teristics of the immune system. As immune reactionsare temporally very dynamic it might be necessary toconcentrate more on time courses of the investigatedreactions, rather than using endpoint determination,which is a snapshot in time that misses most of theevents occurring (Koellner et al., 2005). An exampleof this can be seen in this study: the duration of post-challenge exposure was optimised in order to measureA. salmonicida-specific antibody production, whichis slow in trout. However, this overlooked rapid re-sponses, such as oxidative burst, that are well knownto respond toA. salmonicidawithin very short periodsof time.

A better understanding of the fish immune systemi tinge pa-r pat-t entw fishh ncef ikelyt howt exuald videa ts ofS sh.

A

theU z

ive burst response in female fish only, but not in msh. In general, female trout seemed to be moreitive to effluent exposure than male fish in the stt hand. It is well established that the immune sysf fishes is modulated by hormones such as corrowth hormone, reproductive hormones and otHarris and Bird, 2000). Sex also modulates the stresponse as reproductively active female fish havebserved to demonstrate enhanced cortisol resphile males show depressed cortisol responses toling stress (Pottinger and Carrick, 2000). The sex dif

erences observed in this study were generally coent with the increased responsiveness of female trotress.

The precipitation of theM. lysodeikticussolutionaused by addition of serum samples from maturale fish, gained at the end of the experiment (a2 weeks), is likely also related to sexual deve

s essential for the interpretation of results, suggesffects of environmental contamination on immuneameters in fish. Identification of mechanisms anderns of action of pollutants in the aquatic environmould facilitate estimation of adverse effects onealth. While the study at hand provides clear evide

or adverse effects of sewage effluents, and most lhe PAHs therein, on the fish immune system andhat response can be modulated by stress and sevelopment, further studies are necessary to probetter assessment of possible toxicological effecTP effluents on overall immune competence in fi

cknowledgements

The study was supported by a travel grant fromniversitatsgesellschaft e.V., University of Konstan

B. Hoeger et al. / Aquatic Toxicology 73 (2005) 79–90 89

and by a scholarship to B. Hoeger from the GermanFederal Environmental Foundation, Osnabruck. Re-search in New Zealand was funded by the Founda-tion for Research Science and Technology and by aRoyal Society of New Zealand Travel Grant. The au-thors would like to thank Dr. Willi Nagl, Dr. RosanneEllis, Megan Finley, Murray Smith, Gunter Kotterbaand Tim Charleson.

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