* 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
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: carvanmj{at}uwm.edu.
Received February 28, 2005; accepted May 6, 2005
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ABSTRACT |
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Key Words: TCDD; bioaccumulation; maternal transfer; zebrafish.
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INTRODUCTION |
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While exposure to TCDD impacts early development of fish, reproductive effects of chronic, sublethal exposure to TCDD constitute a growing concern. Reproductive toxicity can be manifested by alterations in gonad development, reproductive and parental behaviors, as well as offspring survival and recruitment (Peterson et al., 1993; Tanguay et al., 2003
; Theobald et al., 2003
). Several factors can influence the reproductive toxicity of TCDD, including absorption rates, tissue distribution, biotransformation, and excretion. The primary exposure route of TCDD accumulation in adult fish is via food (Batterman et al., 1989
; Cook et al., 1990
; Jones et al., 1993
; Stow and Carpenter, 1994
; Thomann, 1989
). Aquatic invertebrates are relatively unaffected by TCDD (Isensee and Jones, 1975
; West et al., 1996
), allowing them to serve as a source of TCDD for fish. The primary route of exposure for larvae is via maternal transfer during vitellogenesis (Ankley et al., 1989
; Cook et al., 1991; Guiney et al., 1979
; Monteverdi and Di Giulio, 2000
; Niimi, 1983
; Vodicnik and Peterson, 1985
). It is important, then, to understand the tissue distribution and translocation of TCDD to offspring in order to more accurately assess the impact of compounds on reproduction and early development in wild fish populations.
The zebrafish (Danio rerio) is a powerful model organism for investigating the molecular and cellular mechanisms by which environmental chemicals disrupt normal developmental processes in both juvenile and adult organisms, as well as during embryonic development (Carvan, et al., 2005; Hill et al., 2005
; Spitsbergen and Kent, 2003
; Teraoka et al., 2003a
). Strong correlations exist between zebrafish and other model vertebrates including birds and mammals (Braunbeck et al., 1992
; Dave and Xiu, 1991
; Mizell and Romig, 1997
; Neilson et al., 1990
; Van Leeuwen et al., 1990
), indicating that zebrafish can be used to predict toxic responses in other species. While great strides have been made, many questions still remain regarding the molecular mechanisms by which TCDD exerts its reproductive toxic response in fishes. The zebrafish is an ideal system for the genetic dissection of the AHR-signaling pathway (Carvan et al., 2005
; Hill et al., 2005
; Tanguay et al., 2003
), and for investigating the effects of TCDD at the earliest stages of development. Additionally, the zebrafish system provides the opportunity to investigate the molecular mechanism(s) of TCDD reproductive toxicity in the context of the whole organism.
The distribution of TCDD to different tissue types and maternal transfer to offspring plays an important role in tissue-specific responses and toxicity. In order to accurately assess the mechanisms by which TCDD exerts its toxicity, it is important to correlate toxicologic effects with measured TCDD concentrations in tissues. The objectives of this study were to determine the distribution and accumulation of TCDD in selected tissues of adult female zebrafish following dietary exposure to TCDD and to correlate tissue concentrations with toxicologic effects. Endpoints of TCDD toxicity were also assessed in embryos to demonstrate early life-stage toxicity from maternally derived TCDD.
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MATERIALS AND METHODS |
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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 013.5 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 conducted 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 semiquantitatively measure impacts on early embryonic development resulting from maternal transfer of TCDD. Subsets of fertilized eggs were transferred to 24-well plates (n = 10 eggs/well with 12 replicates such that n = 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 04 based on the presence of previously characterized endpoints of TCDD toxicity (0 = normal, 1 = one morphologic anomaly, 2 = two morphologic anomalies, 3 = more than two morphologic anomalies, and 4 = 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 = 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 = 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) x 100]) and condition factor (CF, ratio of weight to length, [(length/weight3) x 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.
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RESULTS |
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DISCUSSION |
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The efficiency by which fish assimilate toxic compounds from their food impacts its bioavailability and, therefore, the toxicity of the compound. The half-life of TCDD in fish ranges from 2 months to 2 years, (Cook et al., 1990), and there exist substantial interspecies differences in both bioaccumulation and 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 is consistent with other studies (Giesy et al., 1999
; Isosaari et al., 2004
; Jones et al., 2001
; Tietge et al., 1998
; Walker et al., 1994
). As in this study, assimilation efficiency is often determined by measuring whole-body concentrations following extended dietary exposures and represents the net uptake and elimination (including biotransformation). While quantitative comparisons of assimilation efficiencies are difficult due to differences in experimental design (e.g., life stage of fish exposed), qualitative comparisons can be made. The net assimilation efficiencies of species that are more sensitive to the toxic effects of TCDD such as rainbow trout (858%, Fisk et al., 1997
; Hawkes and Norris, 1977
; Jones et al., 2001
), Atlantic salmon (43%, Isosaari et al., 2004
), and brook trout (8089%, Nichols et al., 1998
; Tietge et al., 1998
), are lower than that for zebrafish (96% in this study). This suggests that bioavailability alone does not account for species-specific differences in TCDD sensitivity.
Differences in tissue distribution between species may impact sensitivity to TCDD toxicity. In zebrafish, overall net tissue distribution of TCDD was similar across treatment groups, with carcass and digestive tract tissues accumulating 9395% of the TCDD burden and brain and ovarian tissues accumulating 46%. It is interesting that, in zebrafish, carcass tissues accumulate greater concentrations than digestive tract tissues, which contain the liver. While similar tissue distributions have been shown for adult brook trout and fingerling rainbow trout and perch (e.g., Kleeman et al., 1986a,b
; Muir and Yarechewski, 1988
; Tietge et al., 1998
), in other species such as carp, burbot, pike perch, salmon, and Baltic herring, TCDD tends to accumulate more in the liver compared to muscle, (e.g., Hektoen et al., 1994
; Korhonen et al., 2001
; Wu et al., 2000
, 2001
). The disparity in tissue allocation likely represents differences in tissue lipid deposition and basic energy utilization and storage. However, TCDD has been shown to impede oogenesis, likely causing the oocytes to be reabsorbed (e.g., Tietge et al., 1998
). Since we observed a dose-dependent impact on OSI, a similar process might occur here, resulting in the redistribution of TCDD to other tissues. Diffusion limitations within the digestive tract may restrict 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 partially explains retention of TCDD within the digestive tract tissues. However, since these tissues also contain the liver, lipid content and the presence of binding proteins could also contribute to such high concentrations occurring in these tissues compared to 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 predict ovarian and egg TCDD concentrations based upon whole-body or muscle tissue levels would aid in human health and ecologic risk assessments. TCDD concentrations in zebrafish ovarian tissues 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
). In zebrafish, ovary concentrations reached 54% of carcass tissue concentration, and eggs contained roughly 6% of carcass TCDD concentrations. This is considerably less than that of trout, in which TCDD concentrations in ovary range from 200% to 703% of muscle concentrations (Jones et al., 2001
; Tietge et al., 1998
), and egg concentrations have been estimated by Walker et al. (1994)
to be 43% of muscle TCDD concentrations. This may reflect differences in ovary and muscle lipid concentrations between the two groups of fishes. Despite these differences, these studies suggest that concentrations in skeletal muscle can be used to predict target tissue and egg residues.
The apparent species-specific differences in the allocation of TCDD to various tissue types raise interesting questions regarding the underlying mechanisms by which certain fish species are more sensitive to the reproductive toxic effects of TCDD. For example, brook trout ovary and eggs constitute 61% and 39% total body burden respectively (Tietge et al., 1998), while zebrafish ovary and eggs carried considerably lower total body burden at 46% and 4% total body burden, respectively. However, observed species-specific differences in reproductive toxicity of TCDD may not be solely due to relative bioavailability (as a measure of percent net dietary assimilation). It has also been suggested that differences in the AHR signaling pathway may contribute to the differential sensitivities of fish species (Hahn, 2001
, 2002
; Zodrow et al., 2004
). The relative contribution that differences in TCDD tissue distribution and variations in the components of the AHR pathway have in determining the sensitivity of organisms/species to TCDD is complicated, unclear, and warrants continued study.
An examination of TCDD-induced reproductive impacts, in addition to TCDD-induced early life stage toxicity, is necessary if we are to more accurately assess the effects such compounds have on wild fish populations. Impacts of sublethal exposure to TCDD on reproduction in fish have not been studied extensively; however, several studies suggest that a sublethal exposure to TCDD perturbs gonadal development, ovulation, and survival of offspring. For example, rainbow trout adults and embryos are equally sensitive to the toxic effects of TCDD (Giesy et al., 2002; Walter et al., 2000
). The lowest observed effect level (LOEL) for reproductive impacts were 1.8 ng TCDD/kg female, resulting in reduced OSI, and transfer of as little as 0.3 ng/kg egg impacted survival of eggs and fry (Giesy et al., 2002
; Jones et al., 2001
). In lake trout, sublethal exposure to TCDD (resulting in approximately 0.380.50 ng/g in skeletal muscle) impacts oocyte viability when concentrations in oocytes are
0.20 ng/g egg, while maternal transfer of 0.050.15 ng/g results in sac fry mortality (Walker et al., 1994
). Adult female brook trout exposed to TCDD (whole-body concentrations ranged from 0.07 to 1.20 ng/g fish) show no adverse effects on survival, growth, gonad development, or egg production. However, accumulation of 1.2 ng/g TCDD causes a delay in initial spawn as well as reduces egg viability (Tietge et 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 completely suppresses spawning activity, which corresponds with arrested gonad development and oocyte atresia. Unfortunately, the small sample size of this experiment (Wannemacher et al., 1992
) makes reproductive toxicity difficult to evaluate, and a dose-response relationship for TCDD-induced reproductive toxicity could not be determined because levels of TCDD were not measured in females or eggs.
The results presented here confirm that, while adult zebrafish are fairly resistant to the toxic effects of TCDD, chronic exposure to sublethal concentrations adversely affected reproduction as measured by impacts on the ovary, as well as offspring health and survival. Chronic dietary exposure resulting in the accumulation of 1.1136 ng/g fish did not suppress spawning activity; however, the ovary was impacted with an accumulation of as little as 0.6 ng/g fish. Since the observed decrease in breeding condition (OSI) was not likely the result of decreased overall health of the fish, it is possible that these effects are the result of direct impairment of the ovary or of the hypothalamic-pituitary-gonadal axis. While we were not able to establish impacts on egg production in this study, preliminary experiments performed in our laboratories using the same dosing regimen demonstrate that egg production is reduced by at least 50% with an estimated accumulation of 3 ng/g TCDD. Even if overall egg production is not greatly reduced, maternal transfer impacts offspring health and survival following the accumulation of as little as 1.1 ng/g fish. Taken together, this suggests that the LOEL for reproductive toxicity in zebrafish is in the range of 0.6 and 1.1 ng/g fish.
Maternal transfer experiments suggest that relatively low concentrations of TCDD are capable of impairing offspring health and survival, and that maternal factors also have a role in early lifestage toxicity of TCDD. Exposure of fertilized zebrafish eggs to waterborne TCDD does not impact survival of 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, with pericardial 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 of TCDD resulted in the accumulation of 0.0421.2 ng/g egg, and significant increases in early life stage toxicity occurred following accumulation of as little as 0.094 ng/g egg. This is considerably lower than the ED50 (2.2 ng/g) for pericardial edema following waterborne exposure of embryos post-fertilization, and even the presumed LOEL (0.8 ng/g) which is coincident with the onset of pericardial edema (Henry et al., 1997
). And while ELS toxicity was correlated with TCDD levels within the eggs (Fig. 6), the slope of the regression suggests that other factors could contribute to the observed poor larval success. Pollutants can evoke an integrated stress response resulting in the reallocation of energy reserves from growth and reproduction, which in turn can impact gamete quality (Pankhurst et al., 1995
; Wendelaar Bonga, 1997
). Additionally, maternally derived transcripts supplied in fertilized eggs play essential roles in early development, including establishment of primordial germ cells and thyroid axis, as well as regulation of gastrulation and subsequent patterning of the dorsal mesoderm (Hashimoto et al., 2004
; Helde and Grunwald, 1993
; Jones et al., 2002
; Kondo et al., 2002
; Martin et al., 1999
; Shinomiya et al., 2000
). Perhaps TCDD exposure has initiated an ovarian toxic response that alters gamete quality or the normal complement of maternal transcripts provided to the egg that are necessary for normal development.
In conclusion, reproductive success of fish can be significantly impaired even when exposed to concentrations of TCDD that do not induce an acute toxic response, alter spawning activity, or decrease egg production. While zebrafish are considerably less sensitive to the reproductive toxicity of TCDD, they show similar impacts on the ovary and survival and health of offspring as the more sensitive species such as trout. As the genomic resources available for zebrafish make them an ideal model system for investigating the molecular mechanisms by which TCDD impacts early development, they also constitute a powerful system to investigate the molecular mechanisms that underlie the ovarian toxic response and reproductive toxicity.
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ACKNOWLEDGMENTS |
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