Aquatic Toxicology 95 (2009) 133–143

Contents lists available at ScienceDirect

Aquatic Toxicology

journa l homepage: www.e lsev ier .com/ locate /aquatox

Changes in reproductive biomarkers in an endangered fish species (bonytailchub, Gila elegans) exposed to low levels of organic wastewater compounds in acontrolled experiment

David B. Walkera,∗, Nicholas V. Parettib,2, Gail Cordyb,2, Timothy S. Grossc,3, Steven D. Zauggd,4,Edward T. Furlongd,5, Dana W. Kolpine,6, William J. Matter f,7, Jessica Gwinna,1, Dennis McIntoshg,8

a University of Arizona, Environmental Research Laboratory, 2601 E. Airport Drive Tucson, AZ 85706, United Statesb United States Geological Survey, Water Resources Division, 520 N. Park Avenue, Suite 221, Tucson, AZ 85719, United Statesc The University of Florida, College of Veterinary Medicine, Center for Environmental and Human Toxicology, Florida Integrated Science Center, Gainesville, FL 32611, United Statesd United States Geological Survey, National Water Quality Laboratory, PO Box 25046, MS 407 Denver Federal Center, Building #95 Denver, CO 80225-0046, United Statese United States Geological Survey, Water Resources Division, PO Box 1230, Iowa City, IA 52240, United Statesf University of Arizona, School of Natural Resources, 325 Biological Sciences East Bldg., Tucson, AZ 85721, United Statesg Delaware State University, 1200 N. DuPont Hwy., Dover, DE 19901, United States

a r t i c l e i n f o

Article history:Received 26 November 2008Received in revised form 21 August 2009Accepted 28 August 2009

Keywords:Bonytail chub17�-Estradiol11-KetotestosteroneVitellogeninEffluent-dependentHistopathology

a b s t r a c t

In arid regions of the southwestern United States, municipal wastewater treatment plants commonlydischarge treated effluent directly into streams that would otherwise be dry most of the year. A betterunderstanding is needed of how effluent-dependent waters (EDWs) differ from more natural aquaticecosystems and the ecological effect of low levels of environmentally persistent organic wastewatercompounds (OWCs) with distance from the pollutant source. In a controlled experiment, we found26 compounds common to municipal effluent in treatment raceways all at concentrations <1.0 �g/L.Male bonytail chub (Gila elegans) in tanks containing municipal effluent had significantly lower levels of11-ketotestosterone (p = 0.021) yet higher levels of 17�-estradiol (p = 0.002) and vitellogenin (p = 0.036)compared to control male fish. Female bonytail chub in treatment tanks had significantly lower con-centrations of 17�-estradiol than control females (p = 0.001). The normally inverse relationship betweenprimary male and female sex hormones, expected in un-impaired fish, was greatly decreased in treat-ment (r = 0.00) versus control (r = −0.66) female fish. We found a similar, but not as significant, trendbetween treatment (r = −0.45) and control (r = −0.82) male fish. Measures of fish condition showed nosignificant differences between male or female fish housed in effluent or clean water. Inter-sex conditiondid not occur and testicular and ovarian cells appeared normal for the respective developmental stage

and we observed no morphological alteration in fish. The population-level impacts of these findings areuncertain. Studies examining the long-term, generational and behavioral effects to aquatic organismschronically exposed to low lev

∗ Corresponding author. Tel.: +1 520 275 7110.E-mail addresses: dwalker@ag.arizona.edu (D.B. Walker), nparetti@usgs.gov

(N.V. Paretti), gcordy@usgs.gov (G. Cordy), tsgross@ufl.edu (T.S. Gross),sdzaugg@usgs.gov (S.D. Zaugg), efurlong@usgs.gov (E.T. Furlong),dwkolpin@usgs.gov (D.W. Kolpin), wmatter@ag.arizona.edu (W.J. Matter),jeg1@email.aizona.edu (J. Gwinn), dmcintosh@desu.edu (D. McIntosh).

1 Tel.: +1 520 626 3322.2 Tel.: +1 520 670 6671.3 Tel.: +1 352 378 8181.4 Tel.: +1 303 236 3269.5 Tel.: +1 303 236 3941.6 Tel.: +1 319 358 3614.7 Tel.: +1 520 621 7280.8 Tel.: +1 302 857 6456.

0166-445X/$ – see front matter © 2009 Elsevier B.V. All rights reserved.doi:10.1016/j.aquatox.2009.08.008

els of OWC mixtures are needed.© 2009 Elsevier B.V. All rights reserved.

1. Introduction

Generally, teleost reproduction is regulated via thehypothalamus–pituitary–gonadal axis, however, the externalenvironmental cues for release by the brain of specific neuro-hormones and neurotransmitters is an evolved, species-specificresponse initiated by seasonal trends in environmental factorssuch as water temperature, photoperiod, electrolyte concen-tration, lunar cycles, and levels of pH and dissolved oxygen. To

varying extents, these external environmental cues are habitat andecosystem-specific.

In light of the many external cues for neurohormone and neuro-transmitter release by the fish brain, it is difficult to isolate the effectof organic wastewater compounds (OWCs) on hormone and protein

1 Toxic

lfiid

mpiipMswswmt(ettaOeoaetr

Oaoircttcibttelp

oogBtNna

m((aiHDao

or effluent added only to make up for evaporative loss and briefback-flushing of filters performed every 2 weeks. This resulted in arenewal rate of approximately 1/3 of each raceways total volumeeach month. Filtration consisted of 250-L sump tanks filled with

Table 1Additive exposure period of bonytail chub to increasing concentrations ofwastewater.

Exposure number Exposure period Percent Percent clean

34 D.B. Walker et al. / Aquatic

evels without excluding external environmental cues responsibleor hormonal fluctuations. Controlled studies of low doses of OWCsn a mixture are relatively rare. Observational field studies are lim-ted in their ability to quantify hormonal effects caused by OWCsue to inherent variability among reference areas.

In most rivers, OWCs are often found at low levels within aatrix of compounds the ecological implications of which are

oorly understood. For example, Kolpin et al. (2002) found dur-ng a national reconnaissance of pharmaceuticals and other OWCsn U.S. streams that of the 30 most frequently detected com-ounds, only about 5% of the concentrations exceeded 1 �g/L.ore recently, Focazio et al. (2008) in another national reconnais-

ance of U.S. streams found that of 63 OWCs detected all but 8ere found at concentrations below 1 �g/L. Snyder (2008) in a

tudy examining OWCs and pharmaceuticals collected from rawater at 20 drinking water plants in the U.S. found much loweredian concentrations of 17 selected compounds ranging from 0.4

o 175 ng/L (0.004–0.175 �g/L, respectively). Kuch and Ballschmiter2001) examining endocrine-disrupting phenolic compounds andstrogens in German rivers found concentrations ranging from 0.9o 199 ng/L (0.009–0.199 �g/L, respectively). Although concentra-ions of OWCs in rivers seem to vary considerably it appears thatll but the most heavily impacted areas will contain mixtures ofWCs with no individual compound exceeding 1 �g/L. The chronicffects of low doses of OWCs in a mixture may be more harmfulver the long-term to individuals or populations than are effects ofcute exposure to a single compound yet studies examining chronicffects of OWC mixtures at environmentally relevant concentra-ions on fish physiology in a controlled environment are relativelyare.

Kolpin et al. (2004) in a study examining urban contribution ofWCs to streams in Iowa found significantly higher concentrationsnd number of compounds detected during low rather than normalr high stream flow conditions. This suggests dilution and flush-ng might have a significant role in overall OWC concentration ofeceiving streams. Effluent-dependent streams in arid regions oftenontain 100% treated wastewater for the majority of the year dueo relatively little dilution or flushing compared to streams in moreemperate areas. As sources of freshwater become scarce, espe-ially in arid regions, water re-use will likely become increasinglymportant to support population growth. This may result in OWCsecoming ubiquitous in freshwater systems previously thoughto be relatively free of these contaminants. Several of these sys-ems currently harbor or are being examined for repatriation ofndangered fish species once native to the region. The effect of lowevel, chronic exposure to OWCs on these endangered species is notresently known.

The Colorado River Basin once contained the greatest numberf endemic fishes of any drainage in the United States. Today, 85%f endemic fish species in Arizona are either threatened or endan-ered with extinction primarily due to habitat loss (Warren andurr, 1994). Given the explosive population growth of humans inhe Lower Colorado River Basin states of Arizona, California, andevada, the effect of OWCs contained within municipal effluent onative fish species, especially those endangered with extinction, istimely and important topic.

Species such as fathead minnow (Pimephales promelas), com-on carp (Cyprinus carpio), guppies (Poecilia reticulate), tilapia

Tilapia mossambica), mosquito fish (Gambusia affinis), medakaOryzias latipes), and goldfish (Carassius auratus) have been useds test organisms to determine the effect of OWCs on teleost phys-

ology and reproduction (Purdom et al., 1994; Bayley et al., 1999;arries et al., 2000; Hutchinson et al., 2003; Sepulveda et al., 2004;iniz et al., 2005; Bistodeau et al., 2006). None of these fishesre native to the arid Southwest where species evolved numer-us morphologic and behavioral mechanisms enabling them to

ology 95 (2009) 133–143

survive and reproduce in the dynamic conditions common to theregion.

We chose to use bonytail chub (Gila elegans) as our test species.Bonytail chub are an endangered warm-water cyprinid native tothe Colorado River watershed and epitomize fish species that haveevolved in aquatic ecosystems of arid regions (Yokota et al., 2000;Snyder et al., 2004). Examples of their evolved morphology includea pronounced pre-dorsal hump in adults, dorsally concave skull,relatively thin caudal peduncle but large caudal fin, small eyes, andembedded scales (USFWS, 2002; Snyder et al., 2004). It is believedthat these morphological adaptations aid in maintaining positionnear the streambed during flooding (Minckley, 1973; Minckley andDeMarais, 2000; USFWS, 2002). Bonytail chub are omnivores andeat insects, small fish, and aquatic plants (Minckley, 1973). Theycan attain a length of 60 cm, weigh up to 2 kg and are tolerant of awide range of water quality conditions (Minckley, 1973; Minckleyand DeMarais, 2000).

Our objective was to determine the effects of chronic exposure(1 year total) to a mixture of OWCs at environmentally relevantconcentrations in municipal effluent on reproductive biomarkers(17�-estradiol, 11-ketotestosterone, vitellogenin, condition factor,sexual maturation, gonado-somatic indices) in a fish species nativeto the arid southwestern U.S. in a controlled environment.

2. Materials and methods

In order to determine potential impacts of OWCs on exposed fishwe established a “clean water” control population for comparison.Clean water used in control raceways consisted of tap water treatedby carbon filtration and reverse osmosis. We constructed 2 controland 2 treatment raceways of marine plywood lined with fiberglasssheeting. To isolate potential effects caused by OWCs we minimizedother environmental variables known to cause hormonal fluctua-tions in fish, primarily water temperature and photoperiod. Wevaried the concentration of effluent by initial dilution with cleanwater so that fish in treatment raceways were exposed to the vol-umetric ratio of 1/3, 2/3, and 3/3 (full strength) effluent to cleanwater respectively in an additive exposure. Each exposure inter-val lasted 4 months and was additive with the same fish exposedto all concentrations throughout the experiment. Fish in treatmentraceways were acclimated by initially exposing them to wastew-ater diluted with clean water and increasing the concentration ofwastewater every 4 months for 1 year. We assigned each 4-monthexposure period an identifying number (Table 1).

In order for fish to survive throughout the experiment, weremoved some of the acutely toxic stressors (e.g., relatively highlevels of ammonia and low dissolved oxygen levels) commonlyfound in wastewater. We achieved this by providing aeration, pas-sive filtration, and nitrification. Our experimental system consistedof 4 large, re-circulating raceways with a capacity of 2500 L each.Water in all raceways was 100% re-circulating with clean water

wastewater byvolume

water byvolume

1 December–April 33 672 April–August 67 333 August–December 100 0

Toxic

2fstrTcibc

mWTecagsetbsRea

viazscstccsaptt

D.B. Walker et al. / Aquatic

.5-cm diameter polyethylene spheres which provided surface areaor nitrifying bacteria. Water from sump tanks was pumped to pres-urized sand filters where a portion was either returned directlyo raceways or funneled through gravity-fed, fluidized sand bedeactors for additional nitrification before re-entering raceways.o minimize seasonal fluctuations in hormone levels caused byhanges in photoperiod and water temperature, we set the light-ng at 12 h of dark and light and maintained water temperatureetween 25 and 28 ◦C by either cooling or heating the buildingontaining the raceways.

We collected secondarily treated wastewater used in the treat-ent raceways from just below the outfall for the Roger Roadastewater Treatment Plant (WWTP) in the Santa Cruz River near

ucson, Arizona. Collection of wastewater corresponded to eachxposure period of the experiment. The Roger Road WWTP has aapacity of 155 million liters per day (mld) and treated an aver-ge of 144 mld from 2004 to 2005. The facility treats wastewaterenerated by a population of approximately 419,000 includingeveral municipal and industrial users. It also receives wastes gen-rated by five major hospitals. The Santa Cruz River upstream ofhe outfall for the Roger Road WWTP had 325 days of zero flowetween 01/01/2004 and 01/01/2005. There are few, if any, fishpecies inhabiting the Santa Cruz River downstream of the Rogeroad WWTP. Reasons for this impoverishment are unknown; how-ver, low dissolved oxygen levels and elevated levels of un-ionizedmmonia are possible causes.

For each additive exposure period, we collected wastewater atarious times of the day to account for diel variability in water qual-ty at the WWTP. We collected treated effluent in sterilized, 5200-Lluminum storage tanks that were stored at the University of Ari-ona’s Environmental Research Laboratory. We collected wateramples for OWCs and general chemistries from treatment andontrol raceways at the beginning and end of each 4-month expo-ure period (Fig. 1). We used a Hydrolab Surveyor 4© Datasondeo measure water temperature, dissolved oxygen, pH, and specificonductivity every 2 weeks on all raceways. To minimize sampleontamination, all researchers avoided use of tobacco, caffeine, per-onal care products, and pharmaceuticals prior to sampling. We

ttached a depth-integrated nozzle connected to a fluorocarbon-olymer bottle and fluorocarbon-polymer coated wand and usedhis to collect composite samples from approximately five equidis-ant points along six different cross sections in each raceway.

Fig. 1. Timeline of sampling events and additions of new effluent for raceways hol

ology 95 (2009) 133–143 135

We used fluorocarbon-polymer equipment and bottles to preventsample contamination from compounds contained in plastic mate-rials. We placed samples on ice and brought them to the UnitedStates Geological Survey, Arizona Water Science Center laboratory(USGS/WRD) in Tucson, Arizona, where we split samples using afluorocarbon-polymer cone-splitter. We split unfiltered samplesdirectly into pre-cleaned, baked amber 1 L glass bottles. We useda fluorocarbon-polymer pump and tubing to run split samplesthrough 0.45-�m glass-fiber filters for samples requiring filtration.Duplicate samples were prepared for laboratory replicates or asbackup samples in case of breakage. Bottles were labeled, sealed,placed on ice, and shipped overnight to the U.S. Geological Survey’sNational Water Quality Laboratory (NWQL) in Denver, Colorado.

We implemented quality assurance procedures to reduce thepossibility of contamination in the field or laboratory. Quality-control measures consisted of field and laboratory blank samplesand a set of internal standards analyzed by the NWQL. We createsfield blank samples by passing a volume of organic-free blank waterthrough clean equipment prior to collecting other samples. At theUSGS/WRD laboratory we created a laboratory blank by passing avolume of clean water through the cone-splitter and then pumpingthis through our filter apparatus before processing other samples.For every 10 samples analyzed, NWQL analyzed at least one fortifiedlaboratory spike and one laboratory blank with their equipment.We added a surrogate spike to samples prior to extraction to moni-tor efficiency or recovery of a sample. Any contamination detectedby the laboratory caused us to remove that compound from theanalyses unless the concentration in the field sample was 10 timesgreater than the concentration reported in the laboratory blank.

The laboratory used two analytical methods to determine con-centrations of 83 OWCs in samples. We selected target compoundswithin each method based on criteria developed from knownor suspected toxicity, potential hormonal activity, persistence inthe environment, and quantity of production and human use.Compounds were categorized based on their primary use andeach compound was placed into one of the following groups;detergent metabolites, fire/flame retardants, fragrances/flavors,fuels/polyaromatic hydrocarbons (PAHs), herbicides/insecticides,

household wastewater compounds, non-prescription drugs, plas-ticizers/antioxidants prescription drugs, and steroids (Kolpin etal., 2002). We extracted human prescription and non-prescriptiondrugs and their select metabolites (18 compounds) by a solid-

ding bonytail chub over 12 months (OWC: organic wastewater compounds).

1 Toxic

pomep

dptipe0dctrTssdlcCr

eupttw(r62opaaso(a

avmcittntvpik3aoi

sti

36 D.B. Walker et al. / Aquatic

hase-extraction method and analyzed with a polar reverse-phasectylsilane (C8) HPLC column and high-performance liquid chro-atography (HPLC). We extracted 65 OWCs using liquid-liquid

xtraction and analyzed with capillary-column gas chromatogra-hy/mass spectrometry.

Prior to analysis, mass spectrometry reference spectra wereetermined with calibration solutions to positively identify com-ounds in the sample. We normalized the response, retentionimes, spectra, and ion abundances to the base peak of the cal-bration compound which we then used for quantitation, and ifossible, we used two qualifier ions for confirmation. The properlution time had to match this retention-time window within.1 min (±6 s) for a positive identification. We used internal stan-ards to calculate compound concentrations from 5- to 8-pointalibration curves (0.01–10.0 �g/L). If the calculated concentra-ion was below the lowest point on the calibration curve or theeporting limit the concentration results are reported as estimated.he reporting limit (RL) is equivalent to the lowest concentrationtandard that we could accurately quantify. Any RL represents atatistically determined reference value above which we reportedata values without qualification. These can be adjusted based on

aboratory performance data. We did not dismiss detections andoncentrations below the RL as the results were properly qualified.oncentrations of compounds with confirmed detections below theeporting level or reported as estimated we used in analyses.

The U.S. Fish and Wildlife Service, Dexter National Fish Hatch-ry and Technology Center in New Mexico provided bonytail chubsed in this study. We housed all fish in clean water for 6 monthsrior to any introduction of effluent into treatment raceways. Bony-ail chub were transported from Dexter National Fish Hatcheryo the University of Arizona’s Environmental Research Laboratoryhere fish were weighed, measured (total length), and randomly

using Random Number Generator Pro©) distributed among fouraceways each containing 63 fish (mean biomass per raceway,.3 kg ± 375 g). We randomly chose 2 raceways as treatment andas controls. We chose 10–11 fish at random out of the total pool

f 63 fish in each raceway for venipuncture and analysis of thelasma hormones 17�-estradiol (E2), 11-ketotestosterone (11-KT),nd the egg protein vitellogenin (VTG). We used the same individu-ls for venipuncture during all subsequent exposures. We randomlyelected 5 fish not used for venipuncture in each raceway at the endf each 4-month exposure period for gonadal and hepatic necropsyFig. 1). At the end of the final exposure, we performed necropsy onll remaining fish in each raceway.

We used tricaine methanesulfonate (MS-222) to sedate fishnd inserted passive integrated transponders (PIT) just below theentral epidermis. We obtained morphological and physiologicaleasures from each fish at the end of each exposure period. We

alculated fish condition factor (K) by dividing length by weight ofndividual fish. Each fish was anesthetized with an MS-222 solu-ion and then weighed, measured (total length), and identified byheir PIT tag. A sterile, 1-mL syringe fitted with a 20-gauge, 1.5-ineedle pretreated with sodium heparin (10–20 units/mL) was usedo obtain blood samples from the caudal vasculature of each pre-iously selected fish. After we obtained whole blood, we quicklylaced fish into an aerated recovery tank. Samples were inserted

nto a microtainer tube pretreated with lithium heparin and wereept on ice until centrifugation. We centrifuged whole blood at000 rpm for 35 min and then transferred plasma to cryovial tubesnd stored these at a temperature of −80 ◦C. We shipped samplesn dry ice overnight to the USGS Florida Caribbean Science Center

n Gainesville, FL, for analysis.

We performed gonadal and hepatic necropsies on 5 randomlyelected fish from each raceway every 4 months (Fig. 1). We anaes-hetized fish and made a fatal cut near the gill cover. We usedris scissors to start an incision above the anus. We used operat-

ology 95 (2009) 133–143

ing scissors (dull side) to open the abdominal cavity for collectionof gonads. We weighed samples to determine the gonado-somaticindex (GSI) calculated as gonad weight/total body weight × 100.We fixed whole gonad samples in buffered formalin and sent themto the USGS Florida Caribbean Science Center where they wereembedded in paraffin, sectioned to 5 �m, and stained with hema-toxylin and eosin for histological evaluation. We classified gonadsof female fish according to four stages of sexual maturation; pre-(1), early- (2), mid- (3), and late-vitellogenic (4). Ovaries contain-ing mostly perinuclear oocytes and cortical alveoli oocytes wereclassified as pre-vitellogenic (stage 1). Early vitellogenic (stage 2)ovaries displayed vitellogenic oocytes with some yolk granules fill-ing the cytoplasm. Mid-vitellogenic (stage 3) ovaries containedlarger oocytes in which the cortical alveoli were pushed to theperiphery of the cell. Ovaries containing oocytes that were fusedin a homogenous mass with yolk granules were considered late-vitellogenic (stage 4). We classified male gonads according to fourstages of sexual maturation; pre- (1), early- (2), mid- (3), andlate-spermatogenic (4). Pre-spermatogenic (stage 1) testes wereimmature and undeveloped and contained only spermatogonia.Testes that had thick germinal epithelium and contained primarilyspermatocytes and spermatids we classified as early spermato-genic (stage 2). Mid-spermatogenic testes (stage 3) had a germinalepithelium of moderate thickness and containing spermatocytes,spermatids, and spermatozoa. Testes with a thin epithelium andcontaining primarily mature spermatozoa which are characteristicof full-grown testes we classified as late-spermatogenic (stage 4).

We pooled water chemistry and plasma biomarker data usedfor parametric analyses from both control and treatment repli-cate raceways. We collected water samples from each raceway atthe beginning and end of each exposure to account for changesand degradation of OWC concentrations (Fig. 1). We then used themean OWC concentration for any given exposure period to com-pare concentrations across all exposures by calculating the ratio ofOWC concentration for an individual exposure/OWC concentrationof all exposures. We used the O’Brien, Brown–Forsythe, Levene, andBartlett tests and found significant differences in group variancesfor plasma biomarker data. Therefore, we used the Welch statisticand weighted the means by the reciprocal of the sample variancesof the group means to determine F statistics and p values for para-metric analyses. To compare relationships between biomarkers weused the pooled data for replicate control and treatment racewaysand divided the covariance of the two variables by the productof their respective standard deviations (Pearson product-momentcorrelation coefficient) to obtain r values. We used JMP© 7.0 by theSAS Institute Inc. for all statistical analyses.

3. Results

We did not anticipate the large temporal variability in the con-centration of compounds released from the treatment plant intothe Santa Cruz River. Wastewater collected during the winter andfall, even when heavily diluted with clean water, had higher thanexpected concentrations of organic compounds (Fig. 2). Reasonsfor this are uncertain, however, the population of Tucson increasessubstantially during this period due to out-of-state visitors and sea-sonal residents. The concentrations of organic compounds duringthe 3 exposure periods were not different enough following dilu-tion with clean water to refer to the volumetric ratios (3:1, 2:1, 1:1)as individual treatments.

Bonytail chub received from Dexter National Fish Hatchery werenot sexually dimorphic. Therefore we could not accurately deter-mine gender of fish used for venipuncture and biomarker analysesuntil the end of the project when we performed gonadal necropsyon these individuals. Sample size of males in either control or treat-

D.B. Walker et al. / Aquatic Toxicology 95 (2009) 133–143 137

F ent tap he tot

mntwtept4oplrtfiv

ig. 2. Concentrations of organic wastewater compounds (OWC) detected in treatmrimary type of use and data are the percentage of that category found divided by t

ent raceways for any given 4-month period was low (exposure 1,= 2 control and 5 treatment males; exposures 2 and 3, n = 2 con-

rol and 4 treatment males [1 mortality]). Sample size of femalesas much higher than males (exposure 1, n = 19 control and 15

reatment females; exposure 2, n = 17 control and 16 treatment;xposure 3, n = 18 control and 16 treatment females). Given the dis-arity in sample size between male and female fish in control andreatment raceways, and the low sample size of males for any given-month additive exposure period, parametric comparisons couldnly be performed by pooling all biomarker results for all exposureeriods at the end of the experiment (Fig. 3). Comparisons of mean

evels of biomarkers in males and females in control or treatmentaceways by individual exposure periods are reported only descrip-ively (Table 3). Slightly fluctuating sample sizes among femalesh were primarily due to difficulty in obtaining sufficient sampleolumes for analyses without exsanguination of individuals.

nks during three periods of exposure. Inset shows groupings of compounds by theiral number of OWCs during a given exposure period.

3.1. General water chemistry

Water temperatures between control (26.89 ± 0.60 ◦C) andtreatment tanks (27.10 ± 0.57 ◦C) were not significantly different(Table 2). Mean dissolved oxygen levels were nearly iden-tical between control and treatment tanks (5.27 ± 0.34 and5.28 ± 0.30 mg/L, respectively). Levels of pH differed betweencontrol and treatment raceways but were still similar withmeans of 7.91(±0.13) and 7.25(±0.28) for control and treatmentraceways, respectively. Mean levels of specific conductance dif-fered between control (1042 ± 93 �S/cm) and treatment raceways

(1818 ± 234 �S/cm).

Total alkalinity, chloride, nitrate, total phosphorous, andorthophosphate were significantly higher in treatment comparedto control raceways (Table 2). Significant nitrification occurredwithin all raceways and levels of nitrate were several orders of

138 D.B. Walker et al. / Aquatic Toxic

Fca

mcb

3

e

ig. 3. Differences in steroidal hormone and egg protein concentrations betweenontrol and treatment bonytail chub by gender. Results are the pooled averages forll additive exposure periods combined.

agnitude higher than ammonia. Levels of arsenic, cadmium,hromium, copper, lead, mercury, selenium, and zinc were allelow detectable limits (0.001–0.01 mg/L).

.2. Organic wastewater compounds

We calculated the ratio of OWC concentration for an individualxposure/OWC concentration of all exposures to compare 4-month

ology 95 (2009) 133–143

additive exposure periods. The number of compounds detectedand their relative concentrations varied over time. Exposure 1 had21 detectable compounds (29% total OWC concentration), expo-sure 2 had 18 compounds detected (45% total OWC concentration),and exposure 3 had 11 compounds detected (26% of the totalOWC concentration). Groups of compounds frequently detectedthroughout the study (as a percentage of total OWC concentra-tion) were prescription drugs (25%), plasticizers (23%), detergentmetabolites (14%), non-prescription drugs (12%) and flame retar-dants (11%).

Compounds originating from prescription drugs in exposures2 and 3 accounted for 31% and 32% of the total concentration ofOWCs, but were only 7% of the OWCs in exposure 1 (Fig. 2). Sul-famethoxazole and carbamazepine comprised most of the totalprescription drug concentration in exposure 1 and were the onlyprescription drugs found in exposures 2 and 3. Sulfamethoxa-zole, an antibiotic commonly prescribed to treat urinary tractinfections, had the highest concentration of any individual com-pound in the study (exposure 2, 0.820 �g/L). Approximately halfas much was detected in exposure 3 (0.431 �g/L) and muchless was found in exposure 1 (0.085 �g/L). We found carba-mazepine, an antiepileptic and mood stabilizing drug, in greatestconcentrations during exposures 2 and 3 (0.339 and 0.278 �g/L,respectively).

Plasticizers and antioxidants, as percentages of total OWCs,were almost identical for all exposures (23%, 23%, and 22% forexposures 1, 2, and 3, respectively). The major compound detectedin these groups was 5-methyl-1H-benzotriazole (5-MeBt) whichaccounted for most or all of the plasticizer and antioxidant concen-trations throughout the experiment and had the second highestconcentration of all individual compounds analyzed. Bisphenola, a known endocrine-disrupting compound (Yokota et al., 2000;Sohoni et al., 2001), was present in low concentrations during expo-sures 1 and 2 (0.016 and 0.018 �g/L, respectively). The suspectedendocrine-disrupting compound para-cresol (wood preservative)(Nishihara et al., 2000) was detected in exposure 1 at 0.005 �g/L.

Detergent metabolites made up 32% of the total OWCconcentration during exposure 1. This group contains biodegra-dation products of the alkylphenol polyethoxylates. Nonylphenoldiethoxylate (NPEO2), nonylphenol monoethoxylate (NPEO1), and4-nonylphenol [(para-nonylphenol, total) (4NP)] were the majorcompounds in this group detected during exposure 1 (0.350, 0.128,and 0.303 �g/L, respectively). The only compound in this groupdetected during exposures 2 and 3 was 4NP, at concentrations ofapproximately 0.180 �g/L in both exposures.

Flame retardants accounted for 12%, 8%, and 15% of the totalOWC concentrations in exposures 1, 2, and 3, respectively. Tri(dichloroisopropyl) phosphate (TCPP) and tri (chloroethyl) phos-phate (TCEP) comprised the total flame retardant concentrationsfor exposures 1 and 3 and a majority of exposure 2 (95%) whichincluded tributyl phosphate. During exposure 1 the concentrationof TCPP was higher than TCEP but during exposures 2 and 3 TCEPconcentrations were much greater than TCPP (Fig. 2).

Of the steroids analyzed, the highest concentrations occurredduring exposure 2 (17% of the OWC concentration), and includedthe fecal sterol, 3-beta-coprostanol, and the plant sterols,beta-stigmastnol and beta-sitosterol. We did not detect thesecompounds during exposure 3 and only 3-beta-coprostanol wasdetected during exposure 1.

Fragrances comprised approximately 3% of the total OWC con-centration for all exposures and individual concentrations of mostcompounds in this group were <0.05 �g/L with the exception ofbenzophenone, a fixative used to prevent ultraviolet light from

damaging scents and colors in perfumes and soaps. We found ben-zophenone, a known endocrine-disrupting compound (Schlumpf etal., 2001), at 0.089 and 0.123 �g/L in exposures 2 and 3, respectively.

D.B. Walker et al. / Aquatic Toxicology 95 (2009) 133–143 139

Table 2General water chemistries of control and treatment raceways.

Analyte Raceway Mean Lower 95% Upper 95% Standard error F ratio � value

Total alkalinity (asmg/L CaCO3)

Control 107.25 88.56 125.948.39 21.13 0.0010Treatment 52.71 34.02 71.41

ChlorideControl 55.54 19.96 91.13

15.97 16.99 0.0021Treatment 148.62 113.04 184.21

SulfateControl 186.50 121.07 251.93

29.37 3.82 0.0793Treatment 267.64 202.20 333.07

Total organiccarbon

Control 5.12 2.47 7.771.19 3.50 0.0910Treatment 8.26 5.61 10.91

Dissolved organiccarbon

Control 5.88 2.58 9.191.48 1.36 0.2707Treatment 8.33 5.02 11.64

AmmoniaControl 0.03 −0.01 0.08

0.02 0.003 0.9560Treatment 0.03 −0.01 0.08

NitrateControl 138.35 38.96 237.74

44.61 17.10 0.0020Treatment 399.20 299.81 498.59

Control 0.96 0.57 1.35

3

udto2rodr

wds

bOific

TM

Total phosphorousTreatment 4.34 3.95

OrthophosphateControl 0.83 0.46Treatment 4.16 3.80

.3. Histology and condition factor

There was a wide distribution of fish in various stages of sex-al maturation in control and treatment raceways. We noted littleifference in sexual maturation in female fish between control andreatment raceways during any individual exposure or at the endf the treatment period. During additive exposure periods 1 and, the majority of females in control (n = 7) and treatment (n = 6)aceways were in either stage 1 or 2 of sexual development. Webserved females in either stage 3 or 4 of sexual developmenturing exposure 3 in both treatment (n = 17) and control (n = 18)aceways.

Male fish in control (n = 10) and treatment (n = 13) racewaysere in similar stages (either 1 or 2) of sexual development foroses 1 and 2. In dose 3, a majority of treatment males were intages 2 and 3 whereas all control males were in stage 1.

Pooled average GSI levels were not significantly different

etween control and treatment females (F1,15.925 = 1.132, p = 0.30).ur sample size of males collected for GSI analysis was low

n control (n = 1) and treatment (n = 4) groups. Of the few malesh where GSI data are available, means were similar betweenontrol (1.39) and treatment (1.51) groups. Fish condition fac-

able 3ean biomarker levels by additive exposure period.

Additive exposure period Biomarker

1

17-� Estradiol (pg/mL)

11-Ketotestosterone (pg/mL)

Vitellogenin (mg/mL)

2

17-� Estradiol (pg/mL)

11-Ketotestosterone (pg/mL)

Vitellogenin (mg/mL)

3

17-� Estradiol (pg/mL)

11-Ketotestosterone (pg/mL)

Vitellogenin (mg/mL)

0.17 188.77 <0.00014.72

1.190.16 208.21 <0.00014.53

tor was not significantly different between control and treatmentfemales (F1,0.592 = 0.598, p = 0.44) or control and treatment males(F1,0.429 = 3.076, p = 0.08).

With the exception of one atretic female housed in a treatmentraceway, growth and morphology of all ovarian and testicular cellsappeared normal for the respective developmental stage with nointer-sex condition noticed.

3.4. Fish plasma hormone and vitellogenin concentrations

At the end of exposure 1, male fish in treatment raceways hadlevels of the primary female hormone E2 5.7 times greater thancontrol males (Table 3). Pooled average E2 levels for all exposureperiods indicate that treatment males had levels greater than 2.5times those of controls (F1,16.621 = 14.176, p = 0.0016; Fig. 3). Thepooled average E2 levels in males in treatment raceways werenearly equivalent with females in control raceways (547 and 568

pg/mL, respectively; Fig. 3).

Pooled average levels of the primary male hormone 11-KTfrom all 3 exposure periods were significantly lower in treatmentcompared to control males (F1,7.150 = 4.735, p = 0.021; Fig. 3). Thegreatest difference in 11-KT levels occurred in dose 1 where control

Sex Control Treatment

M 120.5 (±15.5, n = 2) 688.6 (±133.8, n = 5)F 595.9 (±69.3, n = 19) 432.2 (±55.08, n = 15)M 1026.5 (±295.5, n = 2) 380.2 (±94.5, n = 5)F 591.2 (±66.83, n = 19) 578.9 (±70.4, n = 15)M 0.03 (±0.02, n = 2) 0.44 (±0.14, n = 5)F 0.22 (±0.07, n = 19) 0.17 (±0.08, n = 15)

M 176.5 (±31.5, n = 2) 317.3 (±48.0, n = 4)F 476.2 (±71.9, n = 17) 349 (±33.4, n = 16)M 864.5 (±247.5, n = 2) 617.3 (±150, n = 4)F 677.7 (±87.8, n = 17) 545.1 (±75.5, n = 16)M 0.03 (±0.02, n = 2) 0.10 (±0.01, n = 4)F 0.017 (±0.05, n = 19) 0.21 (±0.04, n = 16)

M 355 (±103, n = 2) 601 (±67.6, n = 4)F 625.9 (±74.5, n = 18) 431.8 (±45.7, n = 16)M 571.5 (±248.5, n = 2) 446.5 (±92.3, n = 4)F 510 (±56.4, n = 18) 470.3 (±58.0, n = 16)M 0.22 (±0.20, n = 2) 0.40 (±0.12, n = 4)F 0.15 (±0.04, n = 18) 0.17 (±0.06, n = 16)

1 Toxic

m(

hfieVtPmt

tpafea(

3

rcVEdabtVr

1

Fc

40 D.B. Walker et al. / Aquatic

ales had 2.7 times the concentrations found in treatment males1027 and 380 pg/mL, respectively; Table 3).

Pooled average levels of the egg protein VTG were 3.5 timesigher in males housed in treatment raceways compared to malesh in control raceways (F1,3.686 = 5.265, p = 0.036; Fig. 3). The great-st difference occurred during the initial exposure where the meanTG concentration was 15 times greater in treatment compared

o control males (0.435 and 0.029 mg/mL, respectively; Table 3).ooled average levels of VTG were 1.8 times greater in treatmentales than in control female fish (0.32 and 0.18 mg/mL, respec-

ively; Fig. 3).There was no significant difference between treatment and con-

rol female fish in pooled average levels of 11-KT (F1,98.957 = 1.151,= 0.286) or VTG (F1,97.081 = 0.0003, p = 0.986; Fig. 3). Pooled aver-ge levels of E2, however, were significantly lower in treatmentemales than controls (F1,10.451 = 11.192, p = 0.001; Fig. 3). The great-st differences occurred during exposure 3 with concentrationspproximately 200 pg/mL less in treatment than control female fishTable 3).

.5. Biomarker relationships

For all additive exposures combined, female fish in controlaceways had concentrations of E2 inversely related to the con-entration of 11-KT (r = −0.66); an inverse relationship betweenTG and 11-KT (r = −0.69); and a positive relationship between2 and VTG (r = 0.48; Fig. 4). These relationships were drasticallyifferent in treatment females with the relationship between E2nd 11-KT decoupling to zero (r = 0.00); the inverse relationshipetween 11-KT and VTG approximately half as strong as in con-

rol fish (r = −0.35); and the positive relationship between E2 andTG (r = 0.28) much lower than that found in female fish in controlaceways.

Control males exhibited an inverse relationship between E2 and1-KT (r = −0.82), 11-KT and VTG (r = −0.59), and a positive rela-

ig. 4. Pearson product-moment correlation coefficients of steroidal hormones and vitehub.

ology 95 (2009) 133–143

tionship between E2 and VTG (r = 0.87; Fig. 4). Treatment malesalso exhibited a decoupling of the relationship between E2 and 11-KT (r = −0.44) and between 11-KT and VTG (r = −0.34). The positiverelationship between E2 and VTG was only slightly less in treatmentcompared to control males (r = 0.70).

4. Discussion

Our results indicate that significant changes in 11-KT, E2, andVTG levels in male fish, in addition to lowered E2 levels in females,occurred following exposure to low levels of environmentallypersistent organic compounds commonly found in treated wastew-ater. With the exception of GSI ratios of males, where samplesize was too low for assumptions, and a slightly higher condi-tion factor in treatment compared to control males, there was noobserved effect in fish morphology. Stage of sexual maturity wassimilar between treatment and control fish and both appeared toexhibit normal development over the course of the experiment.We also observed a significant decoupling of the inverse relation-ship between 11-KT and E2 in treatment fish, especially females,with other less noticeable differences occurring between VTG/E2and VTG/11-KT.

Non-unimodal responses of physiological effect to mixturesof OWCs are well known (Bigsby et al., 1999; Almstrup et al.,2002) thus making relationships between compounds found intreatment raceways to the effects we observed in treatment fishdifficult to establish. Within treatment raceways, we found 10individual compounds known or suspected to cause endocrine dis-ruption in fish. No compound exceeded a concentration of 1.0 �g/L,yet these low levels are environmentally relevant especially for

endangered species where potential repatriation sites are oftenremote. Cordy et al. (2004), in an experiment using 2.4 m longsoil column extracts to determine persistence of OWCs duringrecharge, found eight compounds that were the most persistentin wastewater they collected; carbamazapine, sulfamethoxazole,

llogenin concentrations between treatment and control male and female bonytail

Toxic

bTdmi(a2Spc(dsfof

tadneA1aaTittashbmuanhnte

pppbrot

ewewifstospi

i

D.B. Walker et al. / Aquatic

enzophenone, 5-MeBt, N,N-diethyltoluamide, tributylphosphate,CPP, and cholesterol. We found six of these eight compoundsuring this study and, overall, compounds believed to be environ-entally persistent were those found in the highest concentrations

n treatment raceways. For example, 5-methyl-1H-benzotriazole5-MeBt), a compound routinely used for de-icing of airplanes, issuspected endocrine-disrupting compound (Schaffner and Giger,004) having a physiological effect on both mammals (Chong andugden, 1993) and fish (Pillard et al., 2001). This compound wasersistent throughout this study having the second highest con-entration of any individual compound found. Chong and Sugden1993) demonstrated that 5-methyl-1H-benzotriazole significantlyecrease testis size of adult male Djungarian hamsters (Phodopusungorus) subcutaneously injected with 10 �g of 5-MeBt everydayor 2 months and Pillard et al. (2001) found that concentrationsf 100 mg/L in water of 5-MeBt was acutely toxic to over 50% ofathead minnows exposed for 96 h.

We found alkylphenol polyethoxylates (APEs), or their degrada-ion products alkylphenols (APs), during all three exposures. Bothre ionic surfactants widely used in commercial and householdetergents. Nonylphenol monoethoxylate (NPEO1), and nonylphe-ol diethoxylate (NPEO2) were detected in the first additivexposure period and 4-nonylphenol (4NP) in all exposures.lkylphenols are moderately hydrophobic (Nimrod and Benson,996) and proven to bioaccumulate and cause acute toxicity towide variety of aquatic organisms including fish (Saarikoski

nd Viluksela, 1982; Granmo et al., 1989; McCarty et al., 1993).hese compounds are ubiquitous and environmentally persistentn aquatic ecosystems (Kolpin et al., 2002). Alkyphenols at concen-rations as low as 4.7–7.9 ng L−1 are known to cause cytotoxicityo primary hepatocytes in rainbow trout (Oncorhynchous mykiss)nd induce estrogenicity and vitellogenesis in males of the samepecies (Thorpe et al., 2001). Harries et al. (2000) found that APsave multiple modes of action in male fathead minnows acting asoth an estrogen and anti-androgen. In this study, male fatheadinnows exhibited VTG induction and reduction of secondary sex-

al characteristics when exposed to 4NP to concentrations as lows 0.65 and 8.1 �g/L, respectively. Reproduction in all fathead min-ows was inhibited entirely when fish were exposed to 48 �g/L origher of 4NP. Given the disparity between concentrations of APseeded to elicit a hormonal or reproductive effect in either rainbowrout or fathead minnows it can be inferred that the physiologicalffects of APs are both concentration and genre specific.

The organophosphate esthers (OPEs), tri(2-chloroethyl) phos-hate (TCEP) and tri(dichloroisopropyl) phosphate (TDCP) wereresent in treatment raceways during all 3 exposures. These com-ounds are added to many products to decrease flammability andrittleness. Tri(2-chloroethyl) phosphate has been shown to be aeproductive toxicant in mice (Gulati et al., 1991). Relative to otherrganophosphate esthers, TDCP and TCEP are extremely resistanto environmental degradation (Kawagoshi et al., 2002).

Controlled studies examining physiological biomarkers in fishxposed to low-level mixtures of compounds commonly found inastewater are relatively rare. Rarer still are controlled studies

xamining both sexes simultaneously. Gender-specific differencesere apparent in this study with a strong estrogenic response noted

n treatment males but decreased estrogenicity noted in treatmentemales both as compared to control populations. Treatment maleshowed a strong vitellogenic response with mean levels of VTG inhese fish significantly greater than those found in control malesr even control females. In contrast, treatment and control females

howed no significant difference in plasma VTG levels. This is sur-rising given the large differences in E2 levels noted in the same

ndividuals.It is broadly recognized that E2 is the primary hormone induc-

ng VTG synthesis within fish; however, androgens have also been

ology 95 (2009) 133–143 141

implicated in VTG induction in rainbow trout (Mori et al., 1998),European eel (Anguilla anguilla) (Peyon et al., 1998), and Japaneseeel (A. japonica) (Kwon and Mugiya, 1994) and progesterone isbelieved to play a role, along with E2, in VTG synthesis in thespotted ray (Torpedo marmorata) (Marina et al., 2008). Exact insti-gators of VTG synthesis may be poorly conserved across the phylaand the relative importance of synergy with other steroidogenic orenzymatic compounds may vary by taxonomic group. For femalebonytail chub, decreased levels of E2 for the length of exposure inour experiment did not result in a subsequent decrease in estrogen-receptor mediated vitellogenesis.

Increased plasma E2 levels in male fish can be caused by eitheranti-androgenous (Makynen et al., 2000; Jensen et al., 2004) orestrogenic compounds (MacLatchy et al., 2003) and both have beenshown to reduce plasma 11-KT levels in fish (Fitzpatrick et al., 1993;MacLatchy et al., 2003; Schultz et al., 2003). Filby et al. (2006)showed that at average concentrations of 35.2 ng/L, flutamide, apotent anti-androgen, and 17�-ethynylestradiol (EE2), a potentestrogen, significantly increased expression of CYP19 genes codingfor aromatase production in fathead minnow. The catabolic conver-sion of E2 from testosterone by both brain and gondal aromataseultimately determines plasma E2 concentrations.

We would expect levels of primary male and female hormones,in undisturbed populations in nature, to have an inverse relation-ship. Variability of the relationship of androgens to estrogens isexpected within populations due to seasonality and changes inhabitat and climate, however, the evolutionarily defined range ofthis ratio is always within some population-specific boundary. Wecould therefore assume that deviations outside of this boundaryare a sign of impairment and have the capability to stunt or alterfertility, fecundity, and population viability.

Pooled averages of all exposure periods revealed that plasmaconcentrations of 11-KT and VTG between treatment and controlfemales were not statistically significant, however, we noticed largedifferences in the relationship of 11-KT to E2 in the same individ-uals. The r value of 11-KT and E2 in control females (−0.66) is tobe expected in an un-impaired population. This is in stark contrastto treatment females where the relationship was essentially zero(r = 0.00). We also noticed a decoupling of the relationship between11-KT and E2 to VTG in females but this decoupling was not assignificant as the relationship between 11-KT to E2.

There was no significant difference in VTG levels between con-trol and treatment female fish, but there was a significant decreasein plasma E2 levels. A possible explanation might be the relation-ship between 11-KT and E2. What may be most important regardingfeminization or masculinization is the relationship between andro-gens and estrogens rather than mean levels of either. In treatmentmales, E2 levels were significantly higher than control males andapproached levels found in control females. In females the relation-ship between E2 and 11-KT, although narrowed to zero in treatmentfish, did not invert. This was not the case in males where the E2 to11-KT ratio did invert compared to controls. Without inversion ofthe relationship between 11-KT and E2 in favor of 11-KT in females,we would not expect to find a significant difference in VTG levelsin female fish. For the period we exposed female fish to OWCs inthis experiment, we did not see any change in estrogen receptormediated vitellogenesis.

Although we observed significant differences in E2, VTG, and11-KT levels in males and E2 levels in females, and differences inthe relationships between these variables in all treatment fish, weobserved little to no morphological differences. With the excep-

tion of dose 3 where treatment males were in later stages of sexualmaturation than control males, there was little difference betweencontrol and treatment fish in stage of sexual development. We arenot diminishing the importance of histological or morphometricmeasures in determining the physiological effects of hormonally

1 Toxic

ahfile

ftpmifissslin

A

ttpoaaBHg

R

A

B

B

B

C

C

D

F

F

F

G

42 D.B. Walker et al. / Aquatic

ctive organic compounds in fish, however, it would appear thatormonal or protein changes, at the time and dosages we exposedsh to, can occur in bonytail chub with little or no gross morpho-

ogical change. This is surprising given the relatively long period ofxposure in this study (1 year).

Temporal variability between the occurrence of external cuesor reproduction, the release of various neuro-endocrine stimula-ory compounds, production of steroidal hormones, and eventualrotein development and sexual differentiation in vertebratesakes determination of the ecological consequences of endocrine

mpairment at the population-level difficult. Reproduction insh is highly dependent upon intricate, poorly understood, andpecies-dependent behavioral cues for territoriality, mating, andpawning. We did not measure fertility, fecundity, or behavioro population-level effects are unknown. Studies examining theong-term, generational and behavioral effects to aquatic organ-sms chronically exposed to low levels of OWC mixtures areeeded.

cknowledgements

This work was supported by State of Arizona funding throughhe Technology and Research Initiative. We thank M. Ulibarri and allhe staff at Dexter National Fish Hatchery and Technology Center forroviding expertise, assistance, and generous in-kind support with-ut which this project would not have been possible. We greatlyppreciate the insight of S. Goodbred, and J. Jenkins as reviewersnd of C. Goforth for editing and graphic work. We also thank L.rooke-McDonough, S. Flint, E. Hirleman, C. King, L. Bymers, and. Zerzghi for dedicating their time, and weekends, as animal care-ivers.

eferences

lmstrup, K., Fernández, M.F., Petersen, J.H., Olea, N., Skakkebæk, N.E., Leffers, H.,2002. Dual effects of phytoestrogens result in U-shaped dose-response curves.Environ. Health Perspect. 110, 743–748.

ayley, M., Nielsen, J.R., Baatrup, E., 1999. Guppy sexual behavior as aneffective biomarker of estrogen mimics. Ecotoxicol. Environ. Saf. 43, 68–73.

igsby, R., Chapin, R.E., Daston, G.P., Davis, B.J., Gorski, J., Gray, E., Howdeshell, K.L.,Zoeller, R.T., vom Saal, F.S., 1999. Evaluating the effects of endocrine disrup-tors on endocrine function during development. Environ. Health Perspect. 107,613–618.

istodeau, T.J., Barber, L.B., Bartell, S.E., Cediel, R.A., Grove, K.J., Klauster-meier, J., Woodard, J.C., Lee, K.E., Schoenfuss, H.L., 2006. Larval exposure toenvironmentally relevant mixtures of alkylphenolethoxylates reduces repro-ductive competence in male fathead minnows. Aquat. Toxicol. 79, 268–277.

hong, N.W.S., Sugden, D., 1993. Antigonadal effects of two novel mela-tonin analogues in adult Djungarian hamsters. J. Pineal Res. 15, 104–107.

ordy, G.E., Duran, N.L., Bouwer, H., Rice, R.C., Furlong, E.T., Zaugg, S.D., Meyer,M.T., Barber, L.B., Kolpin, D.W., 2004. Do pharmaceuticals, pathogens, and otherorganic waste water compounds persist when waste water is used for recharge?Ground Water Monit. Remediation 24, 58–69.

iniz, M.S., Peres, I., Pihan, J.C., 2005. Comparative study of the estrogenic responsesof mirror carp (Cyprinus carpio) exposed to treated municipal sewage efflu-ent (Lisbon) during two periods in different seasons. Sci. Total Environ. 349,129–139.

ilby, A.L., Thorpe, K.L., Tyler, C.R., 2006. Multiple molecular effect path-ways of an environmental oestrogen in fish. J. Mol. Endocrinol. 37, 121–134.

itzpatrick, M.S., Pereira, C.B., Schreck, C.B., 1993. In vitro steroid secretion duringearly development of mono-sex rainbow trout: sex differences, onset of pituitarycontrol, and effects of dietary steroid treatment. Gen. Comp. Endocrinol. 91,199–215.

ocazio, M.J., Kolpin, D.W., Barnes, K.K., Furlong, E.T., Meyer, M.T., Zaugg, S.D., Bar-

ber, L.B., Thurman, M.E., 2008. A notional reconnaissance for pharmaceuticalsand other organic wastewater contaminants in the United States-II. Untreateddrinking water sources. Sci. Total Environ. 402, 201–216.

ranmo, A., Ekelund, R., Magnusson, K., Berggren, M., 1989. Lethal and sublethaltoxicity of 4 nonylphenol to the common mussel (Mytilus edulis L). Environ.Pollut. 59, 115–127.

ology 95 (2009) 133–143

Gulati, D.K., Barnes, L.H., Chapin, R.E., Heindel, J., 1991. Final Report on the Repro-ductive Toxicity of Tris(2-chloroethyl)phosphate Reproduction and FertilityAssessment in Swiss CD-1 mice when Administered via Gavage. NTIS TechnicalReport PB92-129170.

Harries, J.E., Runnalls, T., Hill, E., Harris, C.A., Maddix, S., Sumpter, J.P., Tyler, C.R.,2000. Development of a reproductive performance test for endocrine disruptingchemicals using pair-breeding fathead minnows (Pimephales promelas). Environ.Sci. Technol. 34 (3003), 3011.

Hutchinson, T.H., Yokota, H., Hagino, S., Ozato, K., 2003. Development of fish testsfor endocrine disruptors. Pure Appl. Chem. 75, 2343–2353.

Jensen, K.M., Kahl, M.D., Makynen, E.A., Korte, J.J., Leino, R.L., Butterworth, B.C., Ank-ley, G.T., 2004. Characterization of responses to the antiandrogen flutamide ina short-term reproduction assay with the fathead minnow. Aquat. Toxicol. 70,99–110.

Kawagoshi, Y., Nakamura, S., Fukunaga, I., 2002. Degradation of oranophosphoricester in leachate from a sea-based solid waste disposal site. Chemosphere 48,219–225.

Kolpin, D.W., Furlong, E.T., Meyer, M.T., Thurman, E.M., Zaugg, S.D., Barber, L.B.,Buxton, H.T., 2002. Parmaceuticals, hormones, and other organic wastewatercontaminants in U.S. streams 1999 2000. A national reconnaissance. Environ.Sci. Technol. 36, 1202–1211.

Kolpin, D.W., Skopec, M., Meyer, M.T., Furlong, E.T., Zaugg, S.D., 2004. Urbancontribution of pharmaceuticals and other organic wastewater contaminantsto streams during differing flow conditions. Sci. Total Environ. 328, 119–130.

Kuch, H.M., Ballschmiter, K., 2001. Determination of endocrine-distrupting phenoliccompounds and estrogens in surface and drinking water by BRGC-(NCI)-MS inthe pictogram per liter range. Environ. Sci. Technol. 35, 3201–3206.

Kwon, H.C., Mugiya, Y., 1994. Involvement of growth hormone and prolactin in theinduction of vitellogenin synthesis in primary hepatocyte culture of the eel,Anguilla japonica. Gen. Comp. Endocrinol. 93, 51–60.

MacLatchy, D.L., Courtenay, S.C., Rice, C.D., Van der Kraak, G.J., 2003. Development ofa short-term reproductive bioassay using steroid hormone and vitellogenin endpoints in the estuarine mummichog (Fundulus heteroclitus). Environ. Toxicol.Chem. 22, 996–1008.

Makynen, E.A., Kahl, M.D., Jensen, K.M., Tietge, J.E., Wells, K.L., Van der Kraak, G.,Ankley, G.T., 2000. Effects of the mammalian antiandrogen vinclozolin on devel-opment and reproduction of the fathead minnow (Pimephales promelas). Aquat.Toxicol. 48, 461–475.

Marina, P., Salvatore, V., Maria, M.D.F., Fanca, R., Giuseppina, D.G., Maurizio, R., Vin-cenza, L., Ermelinda, L., Piero, A., 2008. Effect of 17�-estradiol and progesteroneon vitellogenesis in the spotted ray Torpedo mamorato Risso 1810 (Elasmo-branchii: Torpediniformes): Studies on females and on estrogen-treated males.Gen. Comp. Endocrinol. 157, 125–132.

McCarty, L.S., Mackay, D., Smith, A.D., Ozburn, G.W., Dixon, D.G., 1993. Residue-basedinterpretation of toxicity and bioconcentration QSARs from aquatic bioassays:polar narcotic organics. Ecotoxicol. Environ. Saf. 25, 253–270.

Minckley, W.L., 1973. Fishes of Arizona. Sims Printing Company Inc., Phoenix, AZ, p.293.

Minckley, W.L., DeMarais, B.D., 2000. Taxonomy of chubs (teleostei, cyprinidae,genus Gila) in the American southwest with comments on conservation. Copeia1, 251–256.

Mori, T., Matisumoto, H., Yakota, H., 1998. Androgen-induced vitellogenin geneexpression in primary cultures of rainbow trout hepatocytes. J. Steroid Biochem.Mol. Biol. 67, 133–141.

Nimrod, A.C., Benson, W.H., 1996. Environmental estrogenic effects of alkylphenolsethoxylates. Crit. Rev. Toxicol. 26, 335–336.

Nishihara, T., Nishikawa, J., Kanayama, T., Dakeyama, F., Saito, K., Imagawa, M., Taka-tori, S., Kitagawa, Y., Hori, S., Utsumic, H., 2000. Estrogenic activities of 517chemicals by yeast two hybrid assay. J. Health Sci. 46, 282–298.

Peyon, P., Calvayrac, R., Baloche, S., Burzawa-Gérard, E., 1998. Metabolic studies oneel (Anguilla anguilla L) hepatocytes in primary culture: effect of 17�-estradioland growth hormone. Comp. Biochem. Physiol. 121, 35–44.

Pillard, D.A., Cornell, J.S., Dufresne, D.L., Mernandez, M.T., 2001. Toxicity of benzo-triazole and benzotriazole derivative to three aquatic species. Water Res. 35,557–560.

Purdom, C.E., Hardiman, P.A., Bye, V.J., Eno, N.C., Tyler, C.R., Sumpter, J.P., 1994.Estrogenic effects of effluents from sewage treatment works. Chem. Ecol. 8,275–285.

Saarikoski, J., Viluksela, M., 1982. Relation between physicochemical properties ofphenols and their toxicity and accumulation in fish. Ecotoxicol. Environ. Saf. 6,501–512.

Schaffner, C., Giger, W., 2004. Anticorrosive benzotriazoles as contaminants inwastewaters and rivers. CHIMIA 58, 453.

Schlumpf, M., Cotton, B., Conscience, M., Haller, V., Steinmann, B., Lichtensteiger, W.,2001. In vitro and in vivo estrogenicity of UV screens. Environ. Health Perspect.109, 239–244.

Schultz, J.R., Petz, L.N., Nardulli, A.M., 2003. Estrogen receptor � and Sp1 regulateprogesterone receptor gene expression. Mol. Cell. Endocrinol. 201, 165–175.

Sepulveda, M.S., Gallagher, E.P., Weiser, C.M., Gross, T.S., 2004. Reproductive and

biochemical biomarkers in largemouth bass sampled downstream of a pulp andpaper mill in Florida. Ecotoxicol. Environ. Saf. 57, 431–440.

Snyder, S.A., 2008. Occurrence, treatment, and toxicological relevance of EDCs andpharmaceuticals in water. Ozone: Sci. Eng. 30, 65–69.

Snyder, E.M., Snyder, S.A., Kelly, K.L., Gross, T.S., Villeneuve, D.L., Fitzgerald, S.D., Vil-lalobos, S.A., Giesy, J.P., 2004. Reproductive responses of common carp (Cyprinus

Toxic

S

T

D.B. Walker et al. / Aquatic

carpio) exposed in cages to influent of the Las Vegas Wash in Lake Mead, Nevada,from late winter to early spring. Environ. Sci. Technol. 38, 6385–6395.

ohoni, P., Tyler, C.R., Hurd, K., Caunter, J., Hetheridge, M., Williams, T., Woods, C.,Evans, M., Toy, R., Gargas, M., Sumpter, J.P., 2001. Reproductive effects of long-

term exposure to bisphenol A in the fathead minnow (Pimephales promelas).Environ. Sci. Technol. 35, 2917–2925.

horpe, K.L., Hutchinson, T.H., Hetheridge, M.J., Scholze, M., Sumpter, J.P., Tyler,C.R., 2001. Assessing the biological potency of binary mixtures of environmentalestrogens using vitellogenin induction in juvenile rainbow trout (Onchoryhnchusmykiss). Environ. Sci. Technol. 35, 2476–2481.

ology 95 (2009) 133–143 143

U.S. Fish and Wildlife Service, 2002. Bonytail (Gila elegans) recovery goals:amendment and supplement to the bonytail chub recovery plan. U.S.Fish and Wildlife Service, Mountain Prairie Region (6), Denver, Colorado.http://www.fws.gov/southwest/es/arizona/Documents/SpeciesDocs/Bonytail

Chub/Bonytail-August-02.pdf.

Warren Jr., M.L., Burr, B.M., 1994. Status of freshwater fishes of the United States:Overview of an imperiled fauna. Fisheries 19, 6–18.

Yokota, H., Tsuruda, Y., Maeda, M., Oshima, Y., Tadokoro, H., Nakazono, A., Honjo,T., Kobayashi, K., 2000. Effect of bishpenol a on the early life stage in Japanesemedaka (Oryzias Latipes). Environ. Toxicol. Chem. 19, 1925–1930.