Accumulation, Tissue Distribution, and Maternal Transfer of Dietary 2,3,7,8,-Tetrachlorodibenzo-p-Dioxin: Impacts on Reproductive Success of Zebrafish

Accumulation, Tissue Distribution, and Maternal Transfer ofDietary 2,3,7,8,-Tetrachlorodibenzo-p-Dioxin: Impacts

on Reproductive Success of Zebrafish

Tisha King Heiden,*,† Reinhold J. Hutz,*,† and Michael J. Carvan, III*,1

*Marine and Freshwater Biomedical Sciences Center, University of Wisconsin-Milwaukee Great Lakes WATER Institute, Milwaukee, Wisconsin 53204,

and †Department of Biological Sciences, University of Wisconsin-Milwaukee, Milwaukee, Wisconsin 53211

Received February 28, 2005; accepted May 6, 2005

TCDD (2,3,7,8-tetrachlorodibenzo-p-dioxin) is a reproductive

toxicant and endocrine disruptor in nearly all vertebrates;

however, the mechanisms by which TCDD alters the reproductive

system is not well understood. The zebrafish provides a powerful

vertebrate model system to investigate molecular mechanisms by

which TCDD affects the reproductive system, but little is known

regarding reproductive toxic response of zebrafish following

chronic, sublethal exposure to TCDD. Here we investigate the

accumulation of TCDD in selected tissues of adult female zebra-

fish and maternal transfer to offspring following dietary exposure

to TCDD (0.08–2.16 ng TCDD/fish/day). TCDD accumulated in

tissues of zebrafish in a dose- and time-dependent manner, except

for brain. Chronic dietary exposure resulting in the accumulation

of 1.1–36 ng/g fish did not induce an overt toxic response or

suppress spawning activity. The ovosomatic index was impacted

with an accumulation of as little as 0.6 ng/g fish, and 10% of the

females showed signs of ovarian necrosis following accumulation

of approximately 3 ng/g TCDD. Offspring health was impacted

with an accumulation of as little as 1.1 ng/g female; thus the

lowest observed effect level (LOEL) for reproductive toxicity in

female zebrafish is approximately 0.6–1.1 ng/g fish. Maternal

transfer resulted in the accumulation of 0.094–1.2 ng/g, TCDD,

which was sufficient to induce the typical endpoints of larval

TCDD toxicity, commonly referred to as blue sac syndrome. This

study provides the necessary framework to utilize the zebrafish

model system for further investigations into the molecular mech-

anisms by which TCDD exerts its reproductive toxic responses.

Key Words: TCDD; bioaccumulation; maternal transfer;

zebrafish.

Polychlorinated dibenzodioxins and polycyclic aromatichydrocarbons are two classes of environmental contaminantsknown to adversely affect reproduction and early developmentin fish (Giesy et al., 1999; Peterson et al., 1993; Polandand Knutson, 1982; Walker et al., 1996). The most toxic

of these compounds is generally considered to be 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD), which is a globallydistributed, highly persistent compound with the potential tomodulate several biological processes that impact growth anddevelopment. Juvenile and larval life stages of fish are the mostsusceptible vertebrates to toxic effects of TCDD (Cook et al.,1993; Peterson et al., 1993; Tanguay et al., 2003; Walker andPeterson, 1994). TCDD can disrupt critical developmentalevents (Carney et al., 2004; Hill et al., 2005; Peterson et al.,1993; Spitsbergen and Kent, 2003; Teraoka et al., 2003a;Walker et al., 1996) and can cause cardiovascular dysfunction,edema, hemorrhages, craniofacial malformations, growth ar-rest, and mortality (Antkiewicz et al., 2005; Belair et al., 2001;Bello et al., 2004; Dong et al., 2002; Elonen et al., 1998; Henryet al., 1997; Hill et al., 2004; Mizell and Romig, 1997; Praschet al., 2003; Tanguay et al., 2003; Walker et al., 1996). Basedon current evidence, lake trout and bull trout larvae are themost sensitive, while zebrafish larvae are the least sensitive fishto the developmental toxic effects of TCDD (LC50 0.05 ng/gand 2.5 ng/g, respectively, Cook et al., 2000; Elonen et al.,1998; Henry et al., 1997). In fish and other vertebrates, thetoxic response to TCDD is mediated by the aryl hydrocarbonreceptor (Lusska et al., 1991; Poland and Knutson, 1982; Safe,1986; Whitlock, 1999); however, there exists a considerablerange in sensitivity among different vertebrate species thatcannot be explained solely by differences in aromatic hydro-carbon receptor biology.

While exposure to TCDD impacts early development of fish,reproductive effects of chronic, sublethal exposure to TCDDconstitute a growing concern. Reproductive toxicity can bemanifested by alterations in gonad development, reproductiveand parental behaviors, as well as offspring survival andrecruitment (Peterson et al., 1993; Tanguay et al., 2003;Theobald et al., 2003). Several factors can influence thereproductive toxicity of TCDD, including absorption rates,tissue distribution, biotransformation, and excretion. Theprimary exposure route of TCDD accumulation in adult fishis via food (Batterman et al., 1989; Cook et al., 1990; Joneset al., 1993; Stow and Carpenter, 1994; Thomann, 1989).

1 To whom correspondence should be addressed at Great Lakes WATER

Institute, University of Wisconsin-Milwaukee, 600 E. Greenfield Avenue,

Milwaukee, WI 53204. Fax: (414) 382-1705. E-mail: [email protected].

� The Author 2005. Published by Oxford University Press on behalf of the Society of Toxicology. All rights reserved.For Permissions, please email: [email protected]

TOXICOLOGICAL SCIENCES 87(2), 497–507 (2005)

doi:10.1093/toxsci/kfi201

Advance Access publication May 18, 2005

by guest on February 5, 2016http://toxsci.oxfordjournals.org/

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Aquatic invertebrates are relatively unaffected by TCDD(Isensee and Jones, 1975; West et al., 1996), allowing themto serve as a source of TCDD for fish. The primary route ofexposure for larvae is via maternal transfer during vitellogen-esis (Ankley et al., 1989; Cook et al., 1991; Guiney et al., 1979;Monteverdi and Di Giulio, 2000; Niimi, 1983; Vodicnik andPeterson, 1985). It is important, then, to understand the tissuedistribution and translocation of TCDD to offspring in order tomore accurately assess the impact of compounds on reproduc-tion and early development in wild fish populations.

The zebrafish (Danio rerio) is a powerful model organismfor investigating the molecular and cellular mechanisms bywhich environmental chemicals disrupt normal developmentalprocesses in both juvenile and adult organisms, as well asduring embryonic development (Carvan, et al., 2005; Hillet al., 2005; Spitsbergen and Kent, 2003; Teraoka et al.,2003a). Strong correlations exist between zebrafish and othermodel vertebrates including birds and mammals (Braunbecket al., 1992; Dave and Xiu, 1991; Mizell and Romig, 1997;Neilson et al., 1990; Van Leeuwen et al., 1990), indicating thatzebrafish can be used to predict toxic responses in otherspecies. While great strides have been made, many questionsstill remain regarding the molecular mechanisms by whichTCDD exerts its reproductive toxic response in fishes. Thezebrafish is an ideal system for the genetic dissection of theAHR-signaling pathway (Carvan et al., 2005; Hill et al., 2005;Tanguay et al., 2003), and for investigating the effects ofTCDD at the earliest stages of development. Additionally, thezebrafish system provides the opportunity to investigate themolecular mechanism(s) of TCDD reproductive toxicity inthe context of the whole organism.

The distribution of TCDD to different tissue types andmaternal transfer to offspring plays an important role in tissue-specific responses and toxicity. In order to accurately assess themechanisms by which TCDD exerts its toxicity, it is importantto correlate toxicologic effects with measured TCDD concen-trations in tissues. The objectives of this study were to deter-mine the distribution and accumulation of TCDD in selectedtissues of adult female zebrafish following dietary exposure toTCDD and to correlate tissue concentrations with toxicologiceffects. Endpoints of TCDD toxicity were also assessed inembryos to demonstrate early life-stage toxicity from mater-nally derived TCDD.

MATERIALS AND METHODS

Experimental fish. Adult zebrafish were obtained from the UWM Marine

and Freshwater Biomedical Sciences Center (Milwaukee, WI), and were

maintained at 26–28�C on a 14-hour light and 10-hour dark cycle in a flow-

through buffered, dechlorinated water system. Males (GL, golden leop-

ard, Ekwill Farms) and females (AB or EK, wild type, Ekwill Farms) were

housed separately and acclimated for several weeks prior to the initiation of

experiments. All experimental procedures were approved by the University of

Wisconsin-Milwaukee Animal Care and Use Committee.

Food preparation. The dietary exposure regimen was designed to expose

adult female zebrafish to nonlethal concentrations of TCDD that would yield

TCDD egg concentrations ranging from 0–13.5 ng TCDD/g egg, which are

within the range expected to be toxic to embryos based on the LD50 of 2.5 ng

TCDD/g egg (Elonen et al., 1998; Henry et al., 1997). The nominal TCDD

concentrations were chosen based on the desired target TCDD concentrations

in the eggs following a protocol described by Tietge et al. (1998). Tritium-

labeled TCDD was synthesized and purified to >99% by the manufacturer

(Eagle-Picher, Lenexa, KS, specific activity 47 Ci/mmol). Food yielding a final

concentration of 272 ng 3H-TCDD/g food was prepared following a protocol

described by Fernandez et al. (1998), with modifications. In brief, the 3H-

TCDD stock solution was diluted in acetone, added to trout chow (Zeigler,

Gardner, PA), swirled to ensure homogeneous distribution, and the acetone

evaporated from the food in a fume hood overnight. This food was then mixed

with uncontaminated trout chow to achieve appropriate concentrations, which

were confirmed by liquid scintillation counting to be 0 (acetone only), 10, 40,

100, and 270 ng 3H-TCDD/g food.

Exposure regimen and experimental design. Experiments were con-

ducted in two phases, preexposure (baseline) and exposure, and were initiated

with 24 females in each exposure group. During the preexposure phase, fish

were fed brine shrimp nauplii and trout chow daily for 3 weeks. During the

exposure phase, females were fed (en masse) the trout chow diet containing

0 (acetone only), 10, 40, 100, or 270 ng of 3H-TCDD/g (ppb) food and brine

shrimp nauplii 5 days a week for a period of 4 weeks. Two days a week, fish

were fed uncontaminated trout chow. Fish were fed to satiation and, based on

the average food consumed per fish, received an estimated applied dose of 0.08,

0.32, 0.80, or 2.16 ng TCDD/female/day. Food consumption by individual fish

was not controlled; however, over the course of the experiment, no fish showed

any significant change in weight (within 2 standard deviations of the mean),

suggesting that each fish received a similar dose within treatment groups.

During both phases of the experiment, mortality and general health were

monitored daily, and females were spawned with untreated males weekly. For

the 0, 10, 40, and 100 ppb treatment groups, embryonic development was

monitored through six days post fertilization (dpf) in order to semiquantita-

tively measure impacts on early embryonic development resulting from mater-

nal transfer of TCDD. Subsets of fertilized eggs were transferred to 24-well

plates (n ¼ 10 eggs/well with 12 replicates such that n ¼ 120 eggs/treatment

group) and raised in zebrafish embryo medium (5 mM NaCl, 0.17 mM KCl,

0.33 mM CaCl, 0.33 mM MgSO4) at 28.5�C through 6 dpf. Embryos were

observed daily, and each egg/embryo/larvae was given a score of 0–4 based on

the presence of previously characterized endpoints of TCDD toxicity (0 ¼normal, 1 ¼ one morphologic anomaly, 2 ¼ two morphologic anomalies, 3 ¼more than two morphologic anomalies, and 4 ¼ dead) to establish a cumulative

early life-stage toxicity score (ELS toxicity score) at 6 dpf. Morphological

anomalies observed include yolk sac, pericardial, and cranial edema, cardiac

malformations, uninflated swim bladder, subcutaneous hemorrhage, shortened

jaw, and tail necrosis.

Harvesting of organs. Fish were fasted for 2 days prior to tissue sampling.

Following 5, 10, 15, and 20 days of control or TCDD dietary exposure, a subset

of females (n ¼ 6 per treatment group) was anesthetized by submersion in

0.1 g/l 3-aminobenzoic acid ethyl ester (MS-222, Sigma). Wet weight and total

length were recorded for each fish, and fish were killed by cervical spinal cord

transection. Ovary, brain, digestive tract (including stomach, intestine, liver,

pancreas, and gall bladder) were removed and weighed. The remaining tissues

of the carcass (muscle, kidney, and bone) were combined and weighed. A

subset of eggs (2 h post fertilization) and larvae (7 days post fertilization) were

also collected and weighed for analysis (n ¼ 10, with four replicates). All

tissues and eggs were then stored at �80�C.

Qualitative observations were made on ovarian development (gross

morphologic observations of healthy or necrotic tissue). Ovosomatic index

(OSI, ovary weight as a proportion of total body weight, [(ovary weight/body

weight) 3 100]) and condition factor (CF, ratio of weight to length, [(length/

498 KING HEIDEN, HUTZ, AND CARVAN


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weight3) 3 100]) were calculated for females in the 0, 10, 40, and 100 ppb

treatment groups.

Determination of TCDD concentrations in adult female tissues and

fertilized eggs. Liquid scintillation counting analysis was performed using

a liquid scintillation counter (Packard Tri-Carb A2300, Hewlett Packard

Instruments, 3H counting efficiency 62%). Tissues were processed according

to manufacturer’s instructions (Packard Bioscience). In brief, tissues were

digested with Soluene 350 (Packard) and decolored using 30% hydrogen

peroxide. 1M Tris, pH4, was used to lower pH to reduce background

chemiluminescence, and samples were counted using Bio-safeII cocktail

(Research Products International). Samples were counted for a maximum of

20 min or until a sigma error level of less than 2% was attained, and 3H-TCDD

concentrations in tissues were calculated based on the specific activity of the

parent compound. Tissues from vehicle-exposed females processed for LSC

showed minimal background for all tissues except carcass tissues, resulting

in calculated background levels of chemiluninesence of 18 ± 0.8 pg/g in ovary,

50 ± 5 pg/g in gut tissues, and 82 ± 6 pg/g in eggs. Due to the highly colored

nature of carcass tissues, concentrations were normalized to background

chemiluminescence. Extraction efficiency was >95% as determined using

spike controls.

Whole-body concentrations (total-body burdens) were calculated by taking

the sum of TCDD (ng) in each tissue divided by the wet weight of each fish (g).

Because actual consumption of contaminated food could not be quantified,

cumulative applied dose was calculated from the amount of applied diet, the

analyzed TCDD concentrations in each diet, and the number of fish in each

treatment group. Net dietary assimilation of TCDD was estimated by

determining the percentage of the cumulative applied dose (ng TCDD per

female) present in the whole body of each fish (ng TCDD per fish).

Data analysis. Statistical analysis of the data was performed using Sigma-

Stat software 2.0 and presented as means ± standard error of the mean (SEM).

Two-way analysis of variance (ANOVA) was used to detect treatment-related

effects on CF, OSI, and tissue concentrations of 3H-TCDD with respect to

‘‘dose versus time.’’ Data were evaluated for homogeneity of variance

(homoscedasticity, Levene Median test) and for normality prior to ANOVA.

Pair-wise multiple comparisons were conducted using the Tukey test with

significant differences identified at p < 0.05. Linear regression was performed

to establish relationships between whole-body TCDD concentrations and

estimated applied dose, and between carcass concentrations and concentrations

in ovary and spawned eggs. Linear regression was also used to correlate egg

TCDD concentrations with ELS toxicity.

RESULTS

Accumulation and Tissue Distribution of TCDD

Females in the 40-, 100-, and 270-ppb treatment groupsshowed a significant increase in TCDD concentrations in alltissues over time (Fig. 1). While TCDD accumulated in braintissue, it was not dose dependent. Only brain tissue from

100 ppb 270 ppb10 ppb 40 ppb

Brain

Ovary

TC

DD

co

ncen

tratio

n (n

g/g

)T

CD

D co

nce

ntratio

n (n

g/g

)

Digestive tract

TC

DD

co

ncen

tratio

n (n

g/g

)T

CD

D co

nce

ntratio

n (n

g/g

)

Carcass

FIG. 1. Tissue concentrations of TCDD (ng TCDD/g wet weight) following dietary exposure to 3H-TCDD. Tissues analyzed were ovary, brain, digestive tract

(includes stomach, intestine, liver, pancreas, and gall bladder), and carcass (includes muscle, kidney, and bone). All tissues, except brain, show a dose- and time-

dependent increase in TCDD concentration over time and appear to approach steady-state following 15 days dietary exposure. Only in the highest treatment group

does TCDD accumulate in significantly higher concentrations in brain. Concentrations of TCDD within each tissue following exposure to food containing 40, 100,

and 270 ng TCDD/g significantly increase over time. Females exposed to food containing 10 ng TCDD/g show accumulation of TCDD within each tissue type, but

concentrations do not significantly increase over time.

DISTRIBUTION AND MATERNAL TRANSFER OF TCDD 499


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females in the 270-ppb treatment group showed significantlyhigher concentrations of TCDD (Fig. 1). While females in the10-ppb treatment groups showed accumulation of TCDD in alltissues, concentrations did not significantly increase over time.Across all treatment groups, the greatest concentration ofaccumulated TCDD was found in carcass tissue and the organsof the digestive tract, ranging from 0.74 to 39 ng/g and 1.31 to83 ng/g, respectively. Appreciable amounts of TCDD were alsofound in ovary and brain tissue, ranging across all treatmentgroups from 0.60 to 17 ng/g and 0.54 to 25 ng/g, respectively.Following 20 days dietary exposure, maternal transfer resultedin the accumulation of 0.094–1.3 ng TCDD/g egg (Fig. 2).While TCDD accumulated in eggs of females in the 10-ppbtreatment group, concentrations did not significantly increaseover time. Females in the 40-, 100-, and 270-ppb treatmentgroups showed a dose-dependent increase in maternal transferover time, but concentrations only appeared to approach steadystate in the eggs of females in the highest treatment group.There was no appreciable loss of TCDD from embryonictissues between 2 h post fertilization (hpf) and 7 days postfertilization (dpf) (data not shown).

Total body burdens increased over time in a dose-dependentmanner (p < 0.01) and, following 20 days of dietary exposure,ranged from 1.1 to 36 ng TCDD/g. The percent body burden foreach tissue, calculated as the proportion of total TCDDdetected in each tissue, is listed in Table 1. Tissue allocationof TCDD did not change over time; however, allocation todigestive tract tissues and carcass tissues were variable acrosstreatment groups. Net dietary assimilation of the differenttreatment groups ranged from 82 to 113%, with no significantdifference between treatment groups. This suggests that whole-

body accumulation of TCDD did not affect assimilation fromfood. The mean net dietary assimilation of TCDD was 96 ±4.5% of the applied dose for all fish over the course of theexperiment (n ¼ 96), and there was a positive linear relation-ship between the estimated applied dose and whole-bodyTCDD concentrations (r2 ¼ 0.93, p < 0.001, Fig. 3). Basedupon the concentration of TCDD detected in spawned eggs andaverage egg production, an estimated 4% of total body burdenwas lost as a result of spawning (across all treatment groups).

Relationships between concentrations of TCDD in carcasstissues and ovary or spawned eggs were determined by linearregression (Fig. 4). Concentrations of TCDD in ovary and eggswere 54% and 6% of carcass concentrations, respectively. Therewas a positive correlation between concentration of TCDD incarcass tissues and the concentrations found in ovarian tissues(r2¼ 0.80, p< 0.001) and transfer to eggs (r2¼ 0.87, p< 0.001).

Fish Health and Reproductive Toxicity

Females did not show an overt toxic response following di-etary exposure to TCDD (i.e., fin necrosis, skin discoloration or

100 ppb 270 ppb10 ppb 40 ppb

TC

DD

c

on

ce

ntra

tio

n (n

g/g

)

FIG. 2. Accumulation of TCDD in eggs (ng TCDD/g wet weight)

following maternal dietary exposure to 3H-TCDD. There was a dose- and

time-dependent increase in maternal transfer over time for females in the 40-,

100-, and 270-ppb treatment groups. While exposure to 10 ng TCDD/g food

results in maternal transfer of TCDD, concentrations do not significantly

increase over time. Concentrations only appear to reach steady state in eggs

from females in the highest treatment group.

TABLE 13H-TCDD Accumulated in Each Tissue Type following 20 Days

Dietary Exposure

10 ppb 40 ppb 100 ppb 270 ppb

Ovary 5.3 ± 0.7 6.6 ± 1.7 4.0 ± 0.6 5.7 ± 0.6

Brain 0.9 ± 0.2 0.5 ± 0.1 0.3 ± 0.1 0.6 ± 0.1

Digestive tract 11 ± 1.3a 18 ± 3.0b 25 ± 0.9c 16 ± 0.7b

Carcass 83 ± 1.0 75 ± 4.4* 70 ± 1.1* 78 ± 1.0

Note. Mean percent body burden (± SEM) of 3H-TCDD accumulated in

each tissue type following 20 days dietary exposure. While percent body

burden in ovary and brain is not dose dependent, tissue allocations to digestive

tract and carcass are variable.a,b,cDiffer significantly from all other values within the row.

*Differ significantly from fish treated at 10 ppb but not 270 ppb.

05

1015202530354045

1.00 10.00 100.00 1000.00Estimated Applied Dose (ng/g)

To

tal b

od

y b

urd

en

(n

g/g

)

y = 0.31x + 0.04R2= 0.93

FIG. 3. Semi-log plot showing net dietary assimilation of TCDD by adult

zebrafish based upon linear regression of whole-body TCDD concentrations

and estimated applied dose. A positive linear relationship exists between

estimated applied dose and measured whole-body concentrations, and mean net

dietary assimilation of TCDD was 96 ± 4.5%.



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lesions, uncoordinated fin movements, subcutaneous hemor-rhages, or death). Condition factors were not different betweentreatment groups (Table 2). While exposure to TCDD did notcause an arrest of spawning activity, there was a significantimpact on the ovary. In all treatment groups, OSI was signif-icantly decreased compared to control following 10 daysdietary exposure; however, OSI did not decrease with contin-ued exposure (Table 2). Additionally, 10% of the females fromthe 40- and 100-ppb treatment groups showed gross morpho-logic signs of ovarian necrosis following accumulation of 3 to15 ng/g fish (data not shown).

Early Life Stage Toxicity Due to Maternal Transferof TCDD

Maternal transfer of TCDD resulted in decreased survivaland health of offspring. Survival 24 hpf was impactedfollowing maternal transfer of 0.3 ng TCDD/g egg. Femalesfrom the 40-ppb treatment group showed an 11% decrease inthe percentage of offspring that survived 24 hpf, while femalesfrom the 100-ppb treatment group showed a 12–15 % decrease.As has been shown in waterborne TCDD exposure (Elonenet al., 1998; Henry et al., 1997), typical endpoints of TCDDtoxicity were not apparent until after hatching (~72 hpf) andresembled blue-sac disease (Fig. 5). While TCDD toxicityresulted in the death of some embryos following 72 hpf,survival was not significantly decreased. There was a dose-dependent increase in the cumulative ELS toxicity score overtime, resulting from accumulation of as little as 0.094 ngTCDD/g egg (Table 3). Accumulation of 0.094, 0.417, and0.924 ng/g TCDD resulted in 19, 25, and 35% of larvae thatexhibited one or more signs of TCDD toxicity following20 days of maternal exposure, respectively (Table 4). Linearregression analysis showed a positive correlation between eggTCDD concentration and ELS toxicity score (Fig. 6, r2 ¼ 0.61,p < 0.001).

DISCUSSION

Adult zebrafish are considerably less sensitive to the toxiceffects of TCDD compared with other fish species studied thusfar. Known LD50 for adult fish species range from 3 ng/g(bluegill) to 16 ng/g (carp and yellow perch) (Kleeman et al.,1988b; van der Weiden et al., 1994). While the acute LD50 foradult zebrafish has not been determined, our data demonstratesthat it may be greater than the levels accumulated by zebrafishin this study. In one study, exposure to 70 ng/g TCDD viaabdominal injection inhibits caudal fin regeneration andinduces hepatic toxicity, but the fish (AB strain) did not showovert signs of toxicity (Zodrow et al., 2004; Zodrow andTanguay, 2003), while in another study, adult zebrafish (strainnot reported, likely Tu) show signs of TCDD toxicity following

0.00

0.20

0.40

0.60

0.80

1.00

1.20

1.40

1.60

Eg

g T

CD

D (n

g T

CD

D/g

)

0

5

10

15

20

25

0 10 20 30 40Carcass TCDD (pg TCDD/g)

0 10 20 30 40Carcass TCDD (ng TCDD/g)

Ovary T

CD

D (p

g T

CD

D/g

)

y = 0.45x + 0.18R2= 0.80

y = 0.03x + 0.25R2= 0.87

FIG. 4. Linear regression of TCDD concentrations in ovary and eggs

compared to carcass concentrations in adult zebrafish. There was a positive

correlation between concentration of TCDD in carcass tissues and the

concentrations found in ovary tissues and eggs, suggesting carcass residues

can be used to predict ovary and egg concentrations.

TABLE 2

Mean Condition Factor (CF) and Ovary Somatic Index (OSI) (± SEM) following Dietary Exposure to TCDD

Dietary

exposure

0 ppb 10 ppb 40 ppb 100 ppb

CF OSI CF OSI CF OSI CF OSI

5 days 15.7 ± 1.9 15.0 ± 2.7 15.9 ± 0.8 9.5 ± 1.1 14.6 ± 1.1 11.0 ± 1.8 16.3 ± 1.8 9.9 ± 1.8

10 days 14.2 ± 0.9 13.5 ± 1.1 13.0 ± 1.0 6.7 ± 0.3* 13.6 ± 1.0 7.2 ± 0.2* 13.0 ± 1.0 6.6 ± 0.2*

15 days 12.9 ± 0.5 15.3 ± 1.6 14.3 ± 0.7 8.5 ± 0.7* 12.9 ± 0.6 8.0 ± 0.7* 14.3 ± 1.2 8.8 ± 0.7*

20 days 15.0 ± 0.8 14.2 ± 0.9 12.7 ± 0.5 9.7 ± 0.8* 12.1 ± 0.4 8.8 ± 1.0* 15.0 ± 1.6 7.2 ± 0.6*

*OSIs are significantly lower compared to control (p < 0.05).



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an acute dietary exposure to levels of 5–20 ng TCDD/fish(Wannemacher et al., 1992). Here we show that followingestimated applied doses of 1.6–43 ng/fish, whole-body TCDDconcentration reached 15 and 36 ng/g in the two highesttreatment groups, without inducing an overt toxic response(AB and EK strains). Differences in sensitivity of zebrafishbetween the two published studies and our study may be theresult of the route and duration of exposure, as well as potentialstrain-dependent differences in TCDD sensitivity. Collectively,these studies indicate that the LD50 for adults may be up to 60times greater than that for embryos (2.5 ng TCDD/g), and thatzebrafish are highly resistant to TCDD compared with otherfish species. Additionally, zebrafish approach the sensitivity of

the more sensitive mammalian species such as the guinea pig,and therefore constitute a good model system for extrapolationto mammalian species, as well as fish.

The efficiency by which fish assimilate toxic compoundsfrom their food impacts its bioavailability and, therefore, thetoxicity of the compound. The half-life of TCDD in fish rangesfrom 2 months to 2 years, (Cook et al., 1990), and there existsubstantial interspecies differences in both bioaccumulationand distribution of TCDD (Boening, 1998; Hektoen et al.,1994; Kleeman et al., 1988). Here we show that, in zebrafish,TCDD accumulated within tissues following dietary exposure,and relatively little was eliminated (Figs. 2 and 3), which isconsistent with other studies (Giesy et al., 1999; Isosaari et al.,

FIG. 5. Representative zebrafish larvae (6 dpf) from females fed diets containing TCDD. Larvae exhibited one or more of the following morphological

anomalies: Cranial (ce), pericardial (pe), or yolk sac edema (yse), uninflated swim bladder (*), subcutaneous hemorrhage (not shown), shortened jaw (j), and tail

necrosis (not shown). Not all larvae were affected (see Table 4). A ¼ control, B ¼ 10 ppb diet, C ¼ 40 ppb diet, and D ¼ 100 ppb diet. Scale bar represents 0.01 mm.

TABLE 3

Mean Cumulative Early Life Stage Toxicity Score of Larvae

following Maternal Dietary Exposure to TCDD

Dietary

exposure 0 ppb 10 ppb 40 ppb 100 ppb

baseline 0.47 ± 0.10 0.32 ± 0.09 0.23 ± 0.09 0.23 ± 0.07

5 days 0.37 ± 0.12 0.49 ± 0.12 0.30 ± 0.09 0.46 ± 0.14

10 days 0.50 ± 0.09 0.28 ± 0.09 0.47 ± 0.12 0.95 ± 0.15*

15 days 0.13 ± 0.07 0.37 ± 0.08 0.77 ± 0.15* 0.79 ± 0.10*

20 days 0.30 ± 0.09 0.72 ± 0.03* 0.76 ± 0.11* 0.97 ± 0.13*

Note. Mean cumulative early life stage toxicity score (± SEM) of larvae

(6 dpf).

*ELS scores are significantly higher compared to baseline (p < 0.05).

TABLE 4

Early Life-Stage Toxicity Score following Maternal Dietary

Exposure to TCDD

Early life-stage toxicity score

dose 0 1 2 3 4

0 89 0 0 0.8 10

10 82 0 0 8 11

40 75 3 0 9 13

100 65 8 2 5 20

Note. Percent larvae (6 dpf) within each early life-stage toxicity score

following 20 days maternal dietary exposure to TCDD; n ¼ 120 larvae per

treatment group.



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2004; Jones et al., 2001; Tietge et al., 1998; Walker et al.,1994). As in this study, assimilation efficiency is oftendetermined by measuring whole-body concentrations follow-ing extended dietary exposures and represents the net uptakeand elimination (including biotransformation). While quanti-tative comparisons of assimilation efficiencies are difficult dueto differences in experimental design (e.g., life stage of fishexposed), qualitative comparisons can be made. The netassimilation efficiencies of species that are more sensitive tothe toxic effects of TCDD such as rainbow trout (8–58%, Fisket al., 1997; Hawkes and Norris, 1977; Jones et al., 2001),Atlantic salmon (43%, Isosaari et al., 2004), and brook trout(80–89%, Nichols et al., 1998; Tietge et al., 1998), are lowerthan that for zebrafish (96% in this study). This suggests thatbioavailability alone does not account for species-specificdifferences in TCDD sensitivity.

Differences in tissue distribution between species mayimpact sensitivity to TCDD toxicity. In zebrafish, overall nettissue distribution of TCDD was similar across treatmentgroups, with carcass and digestive tract tissues accumulating93–95% of the TCDD burden and brain and ovarian tissuesaccumulating 4–6%. It is interesting that, in zebrafish, carcasstissues accumulate greater concentrations than digestive tracttissues, which contain the liver. While similar tissue distribu-tions have been shown for adult brook trout and fingerlingrainbow trout and perch (e.g., Kleeman et al., 1986a,b; Muirand Yarechewski, 1988; Tietge et al., 1998), in other speciessuch as carp, burbot, pike perch, salmon, and Baltic herring,TCDD tends to accumulate more in the liver compared tomuscle, (e.g., Hektoen et al., 1994; Korhonen et al., 2001; Wuet al., 2000, 2001). The disparity in tissue allocation likelyrepresents differences in tissue lipid deposition and basicenergy utilization and storage. However, TCDD has beenshown to impede oogenesis, likely causing the oocytes to bereabsorbed (e.g., Tietge et al., 1998). Since we observeda dose-dependent impact on OSI, a similar process mightoccur here, resulting in the redistribution of TCDD to othertissues. Diffusion limitations within the digestive tract mayrestrict dietary assimilation of compounds with log Kow values

greater than 7 (Gobas et al., 1988; Opperhuizen and Sijm,1990). Since TCDD has a log Kow of 6.8, perhaps this partiallyexplains retention of TCDD within the digestive tract tissues.However, since these tissues also contain the liver, lipid contentand the presence of binding proteins could also contribute tosuch high concentrations occurring in these tissues comparedto ovary and brain, as shown in mammals (Diliberto et al.,1999; Poland et al., 1989)

Since the ovary is a major target organ for TCDD toxicity(Gao et al., 2000; Son et al., 1999), the ability to predictovarian and egg TCDD concentrations based upon whole-bodyor muscle tissue levels would aid in human health and ecologicrisk assessments. TCDD concentrations in zebrafish ovariantissues and eggs were highly correlated with carcass tissues(including skeletal muscle), as in brook trout and rainbow trout(Jones et al., 2001; Nichols et al., 1998; Tietge et al., 1998). Inzebrafish, ovary concentrations reached 54% of carcass tissueconcentration, and eggs contained roughly 6% of carcassTCDD concentrations. This is considerably less than that oftrout, in which TCDD concentrations in ovary range from200% to 703% of muscle concentrations (Jones et al., 2001;Tietge et al., 1998), and egg concentrations have beenestimated by Walker et al. (1994) to be 43% of muscle TCDDconcentrations. This may reflect differences in ovary andmuscle lipid concentrations between the two groups of fishes.Despite these differences, these studies suggest that concen-trations in skeletal muscle can be used to predict target tissueand egg residues.

The apparent species-specific differences in the allocation ofTCDD to various tissue types raise interesting questionsregarding the underlying mechanisms by which certain fishspecies are more sensitive to the reproductive toxic effects ofTCDD. For example, brook trout ovary and eggs constitute61% and 39% total body burden respectively (Tietge et al.,1998), while zebrafish ovary and eggs carried considerablylower total body burden at 4–6% and 4% total body burden,respectively. However, observed species-specific differences inreproductive toxicity of TCDD may not be solely due torelative bioavailability (as a measure of percent net dietaryassimilation). It has also been suggested that differences in theAHR signaling pathway may contribute to the differentialsensitivities of fish species (Hahn, 2001, 2002; Zodrow et al.,2004). The relative contribution that differences in TCDDtissue distribution and variations in the components of the AHRpathway have in determining the sensitivity of organisms/species to TCDD is complicated, unclear, and warrantscontinued study.

An examination of TCDD-induced reproductive impacts, inaddition to TCDD-induced early life stage toxicity, is necessaryif we are to more accurately assess the effects such compoundshave on wild fish populations. Impacts of sublethal exposure toTCDD on reproduction in fish have not been studied exten-sively; however, several studies suggest that a sublethal expo-sure to TCDD perturbs gonadal development, ovulation, and

0.00

0.20

0.40

0.60

0.80

1.00

1.20

0.000 0.200 0.400 0.600 0.800 1.000

Mean Egg TCDD (ng/g)

Mean

E

LS

T

oxicity S

co

re

y = 0.78x + 0.37R2= 0.61

FIG. 6. Linear regression of TCDD concentrations in eggs compared to

early life stage toxicity score. Early life-stage toxicity is correlated to egg

TCDD concentration (p < 0.001).



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survival of offspring. For example, rainbow trout adults andembryos are equally sensitive to the toxic effects of TCDD(Giesy et al., 2002; Walter et al., 2000). The lowest observedeffect level (LOEL) for reproductive impacts were 1.8 ngTCDD/kg female, resulting in reduced OSI, and transfer of aslittle as 0.3 ng/kg egg impacted survival of eggs and fry (Giesyet al., 2002; Jones et al., 2001). In lake trout, sublethal expo-sure to TCDD (resulting in approximately 0.38–0.50 ng/g inskeletal muscle) impacts oocyte viability when concentrationsin oocytes are � 0.20 ng/g egg, while maternal transfer of0.05–0.15 ng/g results in sac fry mortality (Walker et al., 1994).Adult female brook trout exposed to TCDD (whole-body con-centrations ranged from 0.07 to 1.20 ng/g fish) show no adverseeffects on survival, growth, gonad development, or egg pro-duction. However, accumulation of 1.2 ng/g TCDD causesa delay in initial spawn as well as reduces egg viability (Tietgeet al., 1998). Wannemacher et al. (1992) show that, in zebrafish,acute dietary exposure of �5 ng TCDD induces a dose-dependent reduction in egg production and completelysuppresses spawning activity, which corresponds with ar-rested gonad development and oocyte atresia. Unfortunately,the small sample size of this experiment (Wannemacheret al., 1992) makes reproductive toxicity difficult to evaluate,and a dose-response relationship for TCDD-induced repro-ductive toxicity could not be determined because levels ofTCDD were not measured in females or eggs.

The results presented here confirm that, while adult zebrafishare fairly resistant to the toxic effects of TCDD, chronicexposure to sublethal concentrations adversely affected re-production as measured by impacts on the ovary, as well asoffspring health and survival. Chronic dietary exposure result-ing in the accumulation of 1.11–36 ng/g fish did not suppressspawning activity; however, the ovary was impacted with anaccumulation of as little as 0.6 ng/g fish. Since the observeddecrease in breeding condition (OSI) was not likely the resultof decreased overall health of the fish, it is possible that theseeffects are the result of direct impairment of the ovary or of thehypothalamic-pituitary-gonadal axis. While we were not ableto establish impacts on egg production in this study, pre-liminary experiments performed in our laboratories using thesame dosing regimen demonstrate that egg production isreduced by at least 50% with an estimated accumulation of3 ng/g TCDD. Even if overall egg production is not greatlyreduced, maternal transfer impacts offspring health and sur-vival following the accumulation of as little as 1.1 ng/g fish.Taken together, this suggests that the LOEL for reproductivetoxicity in zebrafish is in the range of 0.6 and 1.1 ng/g fish.

Maternal transfer experiments suggest that relatively lowconcentrations of TCDD are capable of impairing offspringhealth and survival, and that maternal factors also have a role inearly lifestage toxicity of TCDD. Exposure of fertilizedzebrafish eggs to waterborne TCDD does not impact survivalof eggs past 48 hpf (Elonen et al., 1998; Henry et al., 1997),and several studies suggest that the cardiovascular system is the

major site of action on TCDD developmental toxicity, withpericardial edema being a sensitive endpoint (Cantrell et al.,1996; Dong et al., 2002; Elonen et al., 1998; Henry et al., 1997;Hill et al., 2005; Teraoka et al., 2003b; Walker and Peterson,1994). Here we showed that, over time, maternal transfer ofTCDD resulted in the accumulation of 0.042–1.2 ng/g egg, andsignificant increases in early life stage toxicity occurred fol-lowing accumulation of as little as 0.094 ng/g egg. This is con-siderably lower than the ED50 (2.2 ng/g) for pericardial edemafollowing waterborne exposure of embryos post-fertilization,and even the presumed LOEL (0.8 ng/g) which is coincidentwith the onset of pericardial edema (Henry et al., 1997). Andwhile ELS toxicity was correlated with TCDD levels withinthe eggs (Fig. 6), the slope of the regression suggests thatother factors could contribute to the observed poor larvalsuccess. Pollutants can evoke an integrated stress responseresulting in the reallocation of energy reserves from growthand reproduction, which in turn can impact gamete quality(Pankhurst et al., 1995; Wendelaar Bonga, 1997). Addition-ally, maternally derived transcripts supplied in fertilized eggsplay essential roles in early development, including estab-lishment of primordial germ cells and thyroid axis, as well asregulation of gastrulation and subsequent patterning of thedorsal mesoderm (Hashimoto et al., 2004; Helde andGrunwald, 1993; Jones et al., 2002; Kondo et al., 2002;Martin et al., 1999; Shinomiya et al., 2000). Perhaps TCDDexposure has initiated an ovarian toxic response that alters gam-ete quality or the normal complement of maternal transcriptsprovided to the egg that are necessary for normal development.

In conclusion, reproductive success of fish can be signifi-cantly impaired even when exposed to concentrations of TCDDthat do not induce an acute toxic response, alter spawningactivity, or decrease egg production. While zebrafish areconsiderably less sensitive to the reproductive toxicity ofTCDD, they show similar impacts on the ovary and survivaland health of offspring as the more sensitive species such astrout. As the genomic resources available for zebrafish makethem an ideal model system for investigating the molecularmechanisms by which TCDD impacts early development, theyalso constitute a powerful system to investigate the molecularmechanisms that underlie the ovarian toxic response andreproductive toxicity.

ACKNOWLEDGMENTS

The authors thank Barbara Wimpee for technical and logistical support,

Sharon Daly for use of the liquid scintillation counter, and Russell Cuhel for

providing valuable expertise and assistance with the radionucleotide analyses.

We would also like to thank the reviewers for this manuscript for their insight

and thoughtful comments, which contributed significantly to the final manu-

script. This work was supported in part by a Pilot Project Grant from the

UWM Institute of Environmental Health and the NIEHS MFBS (TKH and

RJH) Center (ES004184), the National Institutes of Health (RJH, ES011569),

and a Shaw Scientist Award (MJC) from the Shaw Fund of the Greater

Milwaukee Foundation. Conflict of interest: none declared.



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Accumulation, Tissue Distribution, and Maternal Transfer of Dietary 2,3,7,8,-Tetrachlorodibenzo-p-Dioxin: Impacts on Reproductive Success of Zebrafish