Accumulation, Tissue Distribution, and Maternal Transfer of Dietary 2,3,7,8,-Tetrachlorodibenzo-p-Dioxin: Impacts on Reproductive Success of
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《毒物学科学杂志》
Marine and Freshwater Biomedical Sciences Center, University of Wisconsin-Milwaukee Great Lakes WATER Institute, Milwaukee, Wisconsin 53204
Department of Biological Sciences, University of Wisconsin-Milwaukee, Milwaukee, Wisconsin 53211
ABSTRACT
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 zebrafish 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 mechanisms by which TCDD exerts its reproductive toxic responses.
Key Words: TCDD; bioaccumulation; maternal transfer; zebrafish.
INTRODUCTION
Polychlorinated dibenzodioxins and polycyclic aromatic hydrocarbons are two classes of environmental contaminants known to adversely affect reproduction and early development in fish (Giesy et al., 1999; Peterson et al., 1993; Poland and 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 globally distributed, highly persistent compound with the potential to modulate several biological processes that impact growth and development. Juvenile and larval life stages of fish are the most susceptible vertebrates to toxic effects of TCDD (Cook et al., 1993; Peterson et al., 1993; Tanguay et al., 2003; Walker and Peterson, 1994). TCDD can disrupt critical developmental events (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 arrest, and mortality (Antkiewicz et al., 2005; Belair et al., 2001; Bello et al., 2004; Dong et al., 2002; Elonen et al., 1998; Henry et al., 1997; Hill et al., 2004; Mizell and Romig, 1997; Prasch et al., 2003; Tanguay et al., 2003; Walker et al., 1996). Based on current evidence, lake trout and bull trout larvae are the most sensitive, while zebrafish larvae are the least sensitive fish to the developmental toxic effects of TCDD (LC50 0.05 ng/g and 2.5 ng/g, respectively, Cook et al., 2000; Elonen et al., 1998; Henry et al., 1997). In fish and other vertebrates, the toxic response to TCDD is mediated by the aryl hydrocarbon receptor (Lusska et al., 1991; Poland and Knutson, 1982; Safe, 1986; Whitlock, 1999); however, there exists a considerable range in sensitivity among different vertebrate species that cannot be explained solely by differences in aromatic hydrocarbon receptor biology.
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.
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 leopard, 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 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,
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 0–4 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.
RESULTS
Accumulation and Tissue Distribution of TCDD
Females in the 40-, 100-, and 270-ppb treatment groups showed a significant increase in TCDD concentrations in all tissues over time (Fig. 1). While TCDD accumulated in brain tissue, it was not dose dependent. Only brain tissue from females in the 270-ppb treatment group showed significantly higher concentrations of TCDD (Fig. 1). While females in the 10-ppb treatment groups showed accumulation of TCDD in all tissues, concentrations did not significantly increase over time. Across all treatment groups, the greatest concentration of accumulated TCDD was found in carcass tissue and the organs of the digestive tract, ranging from 0.74 to 39 ng/g and 1.31 to 83 ng/g, respectively. Appreciable amounts of TCDD were also found in ovary and brain tissue, ranging across all treatment groups from 0.60 to 17 ng/g and 0.54 to 25 ng/g, respectively. Following 20 days dietary exposure, maternal transfer resulted in the accumulation of 0.094–1.3 ng TCDD/g egg (Fig. 2). While TCDD accumulated in eggs of females in the 10-ppb treatment group, concentrations did not significantly increase over time. Females in the 40-, 100-, and 270-ppb treatment groups showed a dose-dependent increase in maternal transfer over time, but concentrations only appeared to approach steady state in the eggs of females in the highest treatment group. There was no appreciable loss of TCDD from embryonic tissues between 2 h post fertilization (hpf) and 7 days post fertilization (dpf) (data not shown).
Total body burdens increased over time in a dose-dependent manner (p < 0.01) and, following 20 days of dietary exposure, ranged from 1.1 to 36 ng TCDD/g. The percent body burden for each tissue, calculated as the proportion of total TCDD detected in each tissue, is listed in Table 1. Tissue allocation of TCDD did not change over time; however, allocation to digestive tract tissues and carcass tissues were variable across treatment groups. Net dietary assimilation of the different treatment groups ranged from 82 to 113%, with no significant difference between treatment groups. This suggests that whole-body accumulation of TCDD did not affect assimilation from food. The mean net dietary assimilation of TCDD was 96 ± 4.5% of the applied dose for all fish over the course of the experiment (n = 96), and there was a positive linear relationship between the estimated applied dose and whole-body TCDD concentrations (r2 = 0.93, p < 0.001, Fig. 3). Based upon the concentration of TCDD detected in spawned eggs and average egg production, an estimated 4% of total body burden was lost as a result of spawning (across all treatment groups).
Relationships between concentrations of TCDD in carcass tissues and ovary or spawned eggs were determined by linear regression (Fig. 4). Concentrations of TCDD in ovary and eggs were 54% and 6% of carcass concentrations, respectively. There was a positive correlation between concentration of TCDD in carcass tissues and the concentrations found in ovarian tissues (r2 = 0.80, p < 0.001) and transfer to eggs (r2 = 0.87, p < 0.001).
Fish Health and Reproductive Toxicity
Females did not show an overt toxic response following dietary exposure to TCDD (i.e., fin necrosis, skin discoloration or lesions, uncoordinated fin movements, subcutaneous hemorrhages, or death). Condition factors were not different between treatment groups (Table 2). While exposure to TCDD did not cause an arrest of spawning activity, there was a significant impact on the ovary. In all treatment groups, OSI was significantly decreased compared to control following 10 days dietary exposure; however, OSI did not decrease with continued exposure (Table 2). Additionally, 10% of the females from the 40- and 100-ppb treatment groups showed gross morphologic signs of ovarian necrosis following accumulation of 3 to 15 ng/g fish (data not shown).
Early Life Stage Toxicity Due to Maternal Transfer of TCDD
Maternal transfer of TCDD resulted in decreased survival and health of offspring. Survival 24 hpf was impacted following maternal transfer of 0.3 ng TCDD/g egg. Females from the 40-ppb treatment group showed an 11% decrease in the percentage of offspring that survived 24 hpf, while females from the 100-ppb treatment group showed a 12–15 % decrease. As has been shown in waterborne TCDD exposure (Elonen et al., 1998; Henry et al., 1997), typical endpoints of TCDD toxicity were not apparent until after hatching (72 hpf) and resembled blue-sac disease (Fig. 5). While TCDD toxicity resulted 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 over time, resulting from accumulation of as little as 0.094 ng TCDD/g egg (Table 3). Accumulation of 0.094, 0.417, and 0.924 ng/g TCDD resulted in 19, 25, and 35% of larvae that exhibited one or more signs of TCDD toxicity following 20 days of maternal exposure, respectively (Table 4). Linear regression analysis showed a positive correlation between egg TCDD concentration and ELS toxicity score (Fig. 6, r2 = 0.61, p < 0.001).
DISCUSSION
Adult zebrafish are considerably less sensitive to the toxic effects of TCDD compared with other fish species studied thus far. 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 for adult zebrafish has not been determined, our data demonstrates that it may be greater than the levels accumulated by zebrafish in this study. In one study, exposure to 70 ng/g TCDD via abdominal injection inhibits caudal fin regeneration and induces hepatic toxicity, but the fish (AB strain) did not show overt signs of toxicity (Zodrow et al., 2004; Zodrow and Tanguay, 2003), while in another study, adult zebrafish (strain not reported, likely Tu) show signs of TCDD toxicity following an acute dietary exposure to levels of 5–20 ng TCDD/fish (Wannemacher et al., 1992). Here we show that following estimated applied doses of 1.6–43 ng/fish, whole-body TCDD concentration reached 15 and 36 ng/g in the two highest treatment groups, without inducing an overt toxic response (AB and EK strains). Differences in sensitivity of zebrafish between the two published studies and our study may be the result of the route and duration of exposure, as well as potential strain-dependent differences in TCDD sensitivity. Collectively, these studies indicate that the LD50 for adults may be up to 60 times greater than that for embryos (2.5 ng TCDD/g), and that zebrafish are highly resistant to TCDD compared with other fish 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 extrapolation to mammalian species, as well as fish.
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 (8–58%, Fisk et 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 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 93–95% of the TCDD burden and brain and ovarian tissues accumulating 4–6%. 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 4–6% 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 0.38–0.50 ng/g in skeletal muscle) impacts oocyte viability when concentrations in oocytes are 0.20 ng/g egg, while maternal transfer of 0.05–0.15 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 1.11–36 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 0.042–1.2 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.
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 manuscript. 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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Department of Biological Sciences, University of Wisconsin-Milwaukee, Milwaukee, Wisconsin 53211
ABSTRACT
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 zebrafish 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 mechanisms by which TCDD exerts its reproductive toxic responses.
Key Words: TCDD; bioaccumulation; maternal transfer; zebrafish.
INTRODUCTION
Polychlorinated dibenzodioxins and polycyclic aromatic hydrocarbons are two classes of environmental contaminants known to adversely affect reproduction and early development in fish (Giesy et al., 1999; Peterson et al., 1993; Poland and 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 globally distributed, highly persistent compound with the potential to modulate several biological processes that impact growth and development. Juvenile and larval life stages of fish are the most susceptible vertebrates to toxic effects of TCDD (Cook et al., 1993; Peterson et al., 1993; Tanguay et al., 2003; Walker and Peterson, 1994). TCDD can disrupt critical developmental events (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 arrest, and mortality (Antkiewicz et al., 2005; Belair et al., 2001; Bello et al., 2004; Dong et al., 2002; Elonen et al., 1998; Henry et al., 1997; Hill et al., 2004; Mizell and Romig, 1997; Prasch et al., 2003; Tanguay et al., 2003; Walker et al., 1996). Based on current evidence, lake trout and bull trout larvae are the most sensitive, while zebrafish larvae are the least sensitive fish to the developmental toxic effects of TCDD (LC50 0.05 ng/g and 2.5 ng/g, respectively, Cook et al., 2000; Elonen et al., 1998; Henry et al., 1997). In fish and other vertebrates, the toxic response to TCDD is mediated by the aryl hydrocarbon receptor (Lusska et al., 1991; Poland and Knutson, 1982; Safe, 1986; Whitlock, 1999); however, there exists a considerable range in sensitivity among different vertebrate species that cannot be explained solely by differences in aromatic hydrocarbon receptor biology.
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.
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 leopard, 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 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,
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 0–4 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.
RESULTS
Accumulation and Tissue Distribution of TCDD
Females in the 40-, 100-, and 270-ppb treatment groups showed a significant increase in TCDD concentrations in all tissues over time (Fig. 1). While TCDD accumulated in brain tissue, it was not dose dependent. Only brain tissue from females in the 270-ppb treatment group showed significantly higher concentrations of TCDD (Fig. 1). While females in the 10-ppb treatment groups showed accumulation of TCDD in all tissues, concentrations did not significantly increase over time. Across all treatment groups, the greatest concentration of accumulated TCDD was found in carcass tissue and the organs of the digestive tract, ranging from 0.74 to 39 ng/g and 1.31 to 83 ng/g, respectively. Appreciable amounts of TCDD were also found in ovary and brain tissue, ranging across all treatment groups from 0.60 to 17 ng/g and 0.54 to 25 ng/g, respectively. Following 20 days dietary exposure, maternal transfer resulted in the accumulation of 0.094–1.3 ng TCDD/g egg (Fig. 2). While TCDD accumulated in eggs of females in the 10-ppb treatment group, concentrations did not significantly increase over time. Females in the 40-, 100-, and 270-ppb treatment groups showed a dose-dependent increase in maternal transfer over time, but concentrations only appeared to approach steady state in the eggs of females in the highest treatment group. There was no appreciable loss of TCDD from embryonic tissues between 2 h post fertilization (hpf) and 7 days post fertilization (dpf) (data not shown).
Total body burdens increased over time in a dose-dependent manner (p < 0.01) and, following 20 days of dietary exposure, ranged from 1.1 to 36 ng TCDD/g. The percent body burden for each tissue, calculated as the proportion of total TCDD detected in each tissue, is listed in Table 1. Tissue allocation of TCDD did not change over time; however, allocation to digestive tract tissues and carcass tissues were variable across treatment groups. Net dietary assimilation of the different treatment groups ranged from 82 to 113%, with no significant difference between treatment groups. This suggests that whole-body accumulation of TCDD did not affect assimilation from food. The mean net dietary assimilation of TCDD was 96 ± 4.5% of the applied dose for all fish over the course of the experiment (n = 96), and there was a positive linear relationship between the estimated applied dose and whole-body TCDD concentrations (r2 = 0.93, p < 0.001, Fig. 3). Based upon the concentration of TCDD detected in spawned eggs and average egg production, an estimated 4% of total body burden was lost as a result of spawning (across all treatment groups).
Relationships between concentrations of TCDD in carcass tissues and ovary or spawned eggs were determined by linear regression (Fig. 4). Concentrations of TCDD in ovary and eggs were 54% and 6% of carcass concentrations, respectively. There was a positive correlation between concentration of TCDD in carcass tissues and the concentrations found in ovarian tissues (r2 = 0.80, p < 0.001) and transfer to eggs (r2 = 0.87, p < 0.001).
Fish Health and Reproductive Toxicity
Females did not show an overt toxic response following dietary exposure to TCDD (i.e., fin necrosis, skin discoloration or lesions, uncoordinated fin movements, subcutaneous hemorrhages, or death). Condition factors were not different between treatment groups (Table 2). While exposure to TCDD did not cause an arrest of spawning activity, there was a significant impact on the ovary. In all treatment groups, OSI was significantly decreased compared to control following 10 days dietary exposure; however, OSI did not decrease with continued exposure (Table 2). Additionally, 10% of the females from the 40- and 100-ppb treatment groups showed gross morphologic signs of ovarian necrosis following accumulation of 3 to 15 ng/g fish (data not shown).
Early Life Stage Toxicity Due to Maternal Transfer of TCDD
Maternal transfer of TCDD resulted in decreased survival and health of offspring. Survival 24 hpf was impacted following maternal transfer of 0.3 ng TCDD/g egg. Females from the 40-ppb treatment group showed an 11% decrease in the percentage of offspring that survived 24 hpf, while females from the 100-ppb treatment group showed a 12–15 % decrease. As has been shown in waterborne TCDD exposure (Elonen et al., 1998; Henry et al., 1997), typical endpoints of TCDD toxicity were not apparent until after hatching (72 hpf) and resembled blue-sac disease (Fig. 5). While TCDD toxicity resulted 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 over time, resulting from accumulation of as little as 0.094 ng TCDD/g egg (Table 3). Accumulation of 0.094, 0.417, and 0.924 ng/g TCDD resulted in 19, 25, and 35% of larvae that exhibited one or more signs of TCDD toxicity following 20 days of maternal exposure, respectively (Table 4). Linear regression analysis showed a positive correlation between egg TCDD concentration and ELS toxicity score (Fig. 6, r2 = 0.61, p < 0.001).
DISCUSSION
Adult zebrafish are considerably less sensitive to the toxic effects of TCDD compared with other fish species studied thus far. 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 for adult zebrafish has not been determined, our data demonstrates that it may be greater than the levels accumulated by zebrafish in this study. In one study, exposure to 70 ng/g TCDD via abdominal injection inhibits caudal fin regeneration and induces hepatic toxicity, but the fish (AB strain) did not show overt signs of toxicity (Zodrow et al., 2004; Zodrow and Tanguay, 2003), while in another study, adult zebrafish (strain not reported, likely Tu) show signs of TCDD toxicity following an acute dietary exposure to levels of 5–20 ng TCDD/fish (Wannemacher et al., 1992). Here we show that following estimated applied doses of 1.6–43 ng/fish, whole-body TCDD concentration reached 15 and 36 ng/g in the two highest treatment groups, without inducing an overt toxic response (AB and EK strains). Differences in sensitivity of zebrafish between the two published studies and our study may be the result of the route and duration of exposure, as well as potential strain-dependent differences in TCDD sensitivity. Collectively, these studies indicate that the LD50 for adults may be up to 60 times greater than that for embryos (2.5 ng TCDD/g), and that zebrafish are highly resistant to TCDD compared with other fish 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 extrapolation to mammalian species, as well as fish.
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 (8–58%, Fisk et 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 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 93–95% of the TCDD burden and brain and ovarian tissues accumulating 4–6%. 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 4–6% 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 0.38–0.50 ng/g in skeletal muscle) impacts oocyte viability when concentrations in oocytes are 0.20 ng/g egg, while maternal transfer of 0.05–0.15 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 1.11–36 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 0.042–1.2 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.
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 manuscript. 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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