Effects of Subchronic Exposure to a Complex Mixture of Persistent Contaminants in Male Rats: Systemic, Immune, and Reproductive Effects

Michael G. Wade*,1, Warren G. Foster{dagger}, Edward V. Younglai{dagger}, Avril McMahon*, Karen Leingartner*, Al Yagminas*, David Blakey*, Michel Fournier{ddagger}, Daniel Desaulniers* and Claude L. Hughes§

* Growth and Development Section, Environmental and Occupational Toxicology Section, Safe Environments Directorate, Health Canada, Environmental Health Centre, Tunney's Pasture, Ottawa, Ontario, Canada K1A 0L2; {dagger} Department of Obstetrics and Gynecology, McMaster University, Health Sciences Centre, Hamilton, Ontario, Canada L8N 3Z5; {ddagger} INRS, Institut Armand Frappier, 245 Blvd. Hymus, Pointe-Claire, Québec, Canada H9R 1G6; and § Department of Obstetrics and Gynecology, Duke University Medical Center, Durham, North Carolina 27710

Received July 19, 2001; accepted October 29, 2001


    ABSTRACT
 TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
 DISCUSSION
 REFERENCES
 
Human populations throughout the world are exposed daily to low levels of environmental contaminants. The consequences of potential interactions of these compounds to human endocrine, reproductive, and immune function remain unknown. The current study examines the effects of subchronic oral exposure to a complex mixture of ubiquitous persistent environmental contaminants that have been quantified in human reproductive tissues. The dosing solution used in this study contained organochlorines (2,3,7,8-tetrachlorodibenzo-p-dioxin [TCDD], polychlorinated biphenyls [PCBs],p,p`-dichlorodiphenoxydichloroethylene [p,p'-DDE],p,p-dichlorodiphenoxytrichloroethane [p,p'-DDT], dieldrin, endosulfan, methoxychlor, hexachlorobenzene, and other chlorinated benzenes, hexachlorocyclohexane, mirex and heptachlor) as well as metals (lead and cadmium). Each chemical was included in the mixture at the minimum risk level (MRL) or tolerable daily intake (TDI) as determined by the U.S. EPA or ATSDR or, for TCDD, at the no observable effect level (NOEL) used to calculate the TDI. Sexually mature male rats were exposed to this complex mixture at 1, 10, 100, and 1000 times the estimated safe levels daily for 70 days. On day 71, all animals were sacrificed and a variety of physiological systems assessed for toxic effects. Evidence of hepatotoxicity was seen in the significant enlargement of the liver in the 1000x group, reduced serum LDH activity (100x), and increased serum cholesterol and protein levels (both 1000x). Hepatic EROD activities were elevated in animals exposed to10x and above. The mixture caused decreased proliferation of splenic T cells at the highest dose and had a biphasic effect on natural killer cell lytic activity with an initial increase in activity at 1x followed by a decrease to below control levels in response to 1000x. No treatment-related effects were seen on bone marrow micronuclei, daily sperm production, serum LH, FSH, or prolactin levels or weights of most organs of the reproductive tract. The weights of the whole epididymis and of the caput epididymis were significantly decreased at 10x and higher doses, although no effect was seen on cauda epididymal weight. The sperm content of the cauda epididymis was increased at the 1x level but not significantly different from control at higher dose levels. A slight, but significant, increase in the relative numbers of spermatids was seen in the animals from the 1000x group with a trend towards reduced proportion of diploid cells at the same dose. Only minor, nondose related changes were seen in parameters related to condensation of chromatin, as determined by flow cytometry, in epididymal sperm. We conclude that the mixture induced effects on the liver and kidney and on general metabolism at high doses but caused only minor effects on immune function, reproductive hormone levels, or general indices of reproductive function measures. These data suggest that additive or synergistic effects of exposure to contaminants resulting in residue levels representative of contemporary human tissue levels are unlikely to result in adverse effects on immune function or reproductive physiology in male rats.

Key Words: organochlorines; dioxin, PCB; spermatogenesis; hepatotoxicity; EROD; PROD; mixture effects.


    INTRODUCTION
 TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
 DISCUSSION
 REFERENCES
 
On a daily basis humans interact with their environments and, as a consequence, are exposed to a broad spectrum of synthesized chemicals present in the food they eat, the air they breath, and the water they drink. A variety of synthetic halogenated organic compounds such as p,p'-dichlorodiphenoxytrichloroethane (p,p'-DDT) and its persistent metabolite p,p`-dichlorodiphenoxydichloroethylene (p,p'-DDE), hexachlorobenzene (HCB), dioxins and furans, polychlorinated biphenyls (PCBs), other organochlorine pesticides, and metals, such as lead and cadmium, have been released into the environment through human activities and are routinely found as contaminants in tissues collected from the human population in Canada and in other nations throughout the world (Davies and Mes, 1987Go; Mes, 1992Go; Mes and Malcolm, 1992Go; Mes and Marchand, 1987Go; Mes et al., 1990Go; Van Hove Holdrinet et al., 1977Go). A number of studies have also demonstrated that many of these same compounds can be quantified in human reproductive fluids including ovarian follicular fluid and semen (Baukloh et al., 1985Go; Foster et al., 2000Go; Jarrell et al., 1993Go; Pauwels et al., 1999Go; Schlebusch et al., 1989Go). Evidence has also been brought forward in the literature that demonstrates that these contaminants can adversely affect endocrine, reproductive, and immune systems in experimental animals and wildlife (Committee on Hormonally Active Agents in the Environment, 1999Go).

Current understanding of the toxicity of these compounds is based primarily on toxicity studies performed on laboratory animals exposed to a single toxic agent. These studies demonstrate that many of these compounds have effects on the reproductive tract, through the alteration of endocrine physiology, the liver, and thyroid physiology (Brouwer et al., 1998Go; Gray et al., 1989Go; Liu et al., 1995Go; Singh and Pandey, 1989Go). The reported effects are seen, in all cases, at relatively high levels of exposure. The human population is ubiquitously exposed to complex mixtures of these contaminants generally at much lower levels of exposure than those routinely examined in animal toxicity studies and the effects of any interactions between such substances on their toxicity is virtually unknown.

To estimate safe levels of exposure for any given compound, for the human population, all available toxicity data for that compound are synthesized into dose rate through a broadly accepted methodology (Barnes and Dourson, 1988Go) which, theoretically, is the highest rate of exposure that can occur over a moderate length of time that will cause no adverse health impacts. These dose estimates, called minimum risk level, reference dose, tolerable daily intake, etc., are promulgated by various public environmental health organizations and are generally used to gauge specific exposures for their safety. While this approach assumes that there is little interaction between chemical substances in their toxic effects or that the degree of any synergistic increase in toxicity will not exceed the safety factors applied, there have been relatively few studies that have tested these assumptions. Further, there have been a number of studies on the toxic effects of exposure to mixtures on general toxicity (Boyd et al., 1990Go; Gyorkos et al., 1985Go; Ito et al., 1995Go) and several studies examining the effects of complex defined mixtures of ground water contaminants on reproduction (Chapin et al., 1989Go; Heindel et al., 1994Go, 1995Go) or immune function (Germolec et al., 1989Go). In the latter study, the effect of a mixture of 25 ground water contaminants was examined on reproduction in mice and rats using a continuous breeding study and only minor effects of the mixture were demonstrated.

In the current study we have examined the toxicity of low doses, relative to doses generally used in animal toxicity experiments, of a complex mixture, containing contaminants that have been demonstrated to be present in human reproductive tissues and fluids, to liver and general physiology and, specifically, to the reproductive and immune systems in the adult male rat.


    MATERIALS AND METHODS
 TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
 DISCUSSION
 REFERENCES
 
Mixture formulation.
The contaminant mixture was formulated to reflect the types of persistent organic and inorganic contaminants to which the human population of Canada could conceivably be exposed (Table 1Go). Inclusion of organochlorine contaminants in the test mixture was based on their detection in 75% or greater of serum samples collected from Canadian women attending fertility clinics that were analyzed for the presence of 45 different organochlorine contaminants (Foster, 1996Go; Jarrell et al., unpublished findings). The dose levels of each component in the mixture reflect the currently promulgated safe levels of exposure as published by the U.S. ATSDR (minimum risk level; MRL), Canadian Environmental Protection Act Chemical Assessments (tolerable daily intake; TDI) or U.S. EPA (reference dose; RfD) or at the lowest NOEL available in the scientific literature (Table 1Go). Neat preparations of mixture components were weighed into a glass bottle, wrapped in tin foil, in amounts appropriate to make the 1000x stock. Other mixture components were first dissolved in ethanol (aldrin, DDT, dieldrin, endosulfan, 1,2,3,4-tetrachlorobenzene) or in diethylether (TCDD) and the appropriate volume added to the stock bottle. Sufficient corn oil to make the appropriate concentration was then added to the bottle, which was heated to 40°C in a sonicating water bath until no undissolved material was visible (approx. 2 h). Lower doses (1x, 10x, 100x) of the dosing solution were prepared by 10-fold serial dilution of the 1000x stock. All dosing solution for the entire experiment was prepared in a single batch and was stored in the dark at 4°C for the duration of the study. The concentration of mixture components in the dosing solution is considered nominal as the actual concentrations of these was not determined through subsequent analyses of the dosing solution.


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TABLE 1 Composition of Contaminant Mixtures Administered to Rats
 
Animals.
Sexually mature Sprague-Dawley rats (45 days of age) were purchased from Charles River (St. Constant, Québec, Canada) and acclimated to holding facilities for one week prior to commencement of dosing. Animals were caged in pairs in clear plastic cages containing wood chips for bedding and maintained under controlled temperature (24°C), humidity (30–70%), and light (12:12 light:dark). All animals were provided standard laboratory rat chow and water ad libitum. Animal care and handling were in accordance with Canadian Council for Animal Care guidelines.

Rats were randomly assigned to control (n = 9), 1x (n = 10), 10x (n = 10), 100x (n = 10), or 1000x (n = 10) treatment groups. Rats received corn oil (vehicle) or appropriate treatment in a volume of 1 µl/g body weight by gavage daily for 70 days, equivalent to one full cycle of spermatogenesis. Animals were monitored daily for signs of overt toxicity and body weight was recorded every 2–3 days.

On the day following the final dose, rats were anesthetized with isofluorane and, after recording body weight, were sacrificed by exsanguination via cardiac puncture followed by decapitation. Blood was collected in Serum Separator Vacutainer tubes (Becton-Dickenson), held for no more than 4 h before serum was collected by centrifugation (3000 x g for 15 min), aliquoted and frozen at –80°C until analysis. Analysis of serum biochemistry was conducted using a Technicon RA-XT (Bayer Diagnostics, Tarrytown, NY) and assay of circulating levels of LH, FSH, and prolactin was conducted using commercial RIA kits (Amersham Pharmacia, Piscataway, NJ). Prior to dissection of the abdominal organs, the body surface was wiped with 70% ethanol, an incision was made into the body cavity and, after the diaphragm was transected, the spleen was removed aseptically and teased into small pieces in complete RPMI-1640 containing 5% fetal calf serum, 100 U/ml penicillin and 100 µg/ml streptomycin. Spleens were kept on ice and shipped to Montréal where splenocyte function was assessed.

The liver, thymus gland, left kidney, pituitary gland, adrenal glands, and major organs of the reproductive tract were removed, dissected free of connective tissue, and weighed. Once weighed, 4 pieces of the liver, roughly 1 g each, were frozen in liquid nitrogen and stored at –80°C prior to extraction of microsomes for hepatic enzyme analysis. An additional piece of liver, taken from the median lobe, was fixed in 4% paraformaldehyde for 72 h for histopathological analysis. One testis from each animal was snap frozen in liquid nitrogen for assessment of daily sperm production and the second testis was placed in ice cold M199 and held on ice until dispersion of testicular cells for flow cytometric analysis. As the amount of fluid lost from seminal vesicles due to dissection varied between animals, seminal vesicle weight was determined after all fluid had been extruded from the structure.

Splenocyte Function
The rationale for the selection of the mitogenic assay and the natural killer activity in this study is the following. Because it was not possible to stimulate an immune response by injection of an antigen (possible effects of the immune response on other physiological parameters studied in the experiment) we had to choose parameters allowing the assessment of the immune status of the animals in relation to the putative toxicity of the mixture. The ability of the mitogenic assay and NK activity to predict an immunotoxic outcome is 67 and 69% respectively when these tests are used alone. However, when they are used in combination, the predictability increases up to 79% (Luster et al., 1992Go). Moreover, in recent studies designed to look at the immunotoxicity of complex mixtures in rat as well as in mouse these immunological markers appeared to be sensitive endpoints to assess the toxic action of chemicals (Fournier et al., 2000Go; Lapierre et al., 1999Go; Omara et al., 1997Go, 1998Go, 2000Go; Tryphonas et al., 1998aGo,bGo).

Splenocytes were put in suspension and washed twice in HBSS with centrifugation. Viable leucocytes were isolated using density gradient centrifugation on Ficoll-paque (Amersham Pharmacia, Piscataway, NJ). The splenocytes were collected from the Ficoll-paque interface and washed 3 times in HBSS. After the final wash, the splenocytes were suspended in complete RPMI-1640. The viability of splenocytes was determined by membrane permeability to propidium iodide using a flow cytometer. The cells were collected and labeled with propidium iodide (1.5 µg/ml) before data acquisition by flow cytometry (FACScan, Becton-Dickinson) to distinguish live and dead cells.

Determination of mitogen-induced lymphocyte blastogenesis.
The ability to undergo blastogenesis was measured by [3H]thymidine uptake. Briefly, splenocytes (5 x 105 per well) were cultured for 72 h at 37°C and 5% CO2 with various concentrations of mitogen in 96-well flat-bottom plates. The optimal concentrations used to stimulate splenocytes were 2.5 µg/ml and 5 µg/ml for concavalin A (Con A), and 2.5 µg/ml and 10 µg/ml for phytohemaglutinin (PHA), both being T-lymphocyte mitogens. After a 72-h incubation, 0.5 µCi of [3H]thymidine was added to each well and plates were incubated for a further 18 h. The cells were harvested onto filters with a Tetratex Cell Harvester, and the amount of radioactivity incorporated was quantitated in a scintillation counter, (Beckman LS1801). The results were expressed as disintegration per minute (dpm).

Determination of NK activity.
To determine the possible effects of the mixture on NK cell activity, splenic NK cell-mediated lysis of murine lymphoma cells (Yac-1 cells) was determined by a flow cytometry method (Brousseau et al., 1998Go; Chang et al., 1993Go). Briefly, target cells (YAC) in exponential phase are counted and 107 cells are incubated with 3–3`-dioctadecycloxacarbocyanine perchlorate (DiO) for 20 min at 37°C with 5% CO2. The cells are then washed and the concentration adjusted at 1 x 106 cells/ml. DiO has an absorption and fluorescence spectra compatible with FITC. Target cells (DiO+) are then incubated with effector cells at different ratios (1/30, 1/60) for 4 h in the presence of propidium iodide (PI). At the end of the incubation period the cells are gently dislodged and data were acquired by FACScan (Becton Dickinson) flow cytometer with a threshold set at FL1 (DIO+ cells) to exclude effector cells and bare cell nuclei that are DiO negative. A total number of 10,000 events are collected. Data were analysed by LYSYS II software (Becton Dickinson) to determine target cell death (cytolethality). When an NK cell damages the membrane of a target cell, PI can no longer be excluded and the target cell is stained by PI to become DiO+, PI+. In contrast, target cells not affected by NK cells will exclude PI to remain DiO+, PI–. The results are expressed in percentage of cytotoxicity.

Liver microsomal enzyme activity.
To assess the activities of several inducible enzymes, a frozen sample of liver was homogenized in 2.5 volume of ice-cold 0.2M Tris containing 1.15% KCl (pH 7.4) with a teflon-glass homogenizer, to prepare a 33% (w/v) homogenate. Nuclei and mitochondria were removed by centrifugation at 10,000 x g, 4°C for 20 min. This postmitochondrial supernatant, (S9) was recovered and stored at –80°C until assayed. Protein concentration was determined by Bio-Rad Bradford protein assay (Bio-Rad, Hercules, CA) using bovine serum albumin (Fraction V, Sigma Chemical, St. Louis, MO) as the standard. In order to determine the effects of the mixture on the induction of hepatic phase I biotransformation monooxygenases, aliquots of S9 were thawed on ice and assayed for ethoxy- and pentoxy-resorufin-O-dealkylase (EROD and PROD, respectively) activities as previously described (Burke et al., 1985Go; Lubet et al., 1985Go) to estimate the activities of CYP1A1 and CYP2B gene products, respectively. These in turn indicate the activation of the aryl hydrocarbon receptor signalling pathway and phenobarbitol-like xenobiotic response. In addition the activity of 7-benzyloxy-resorufin-O-dearylase (BROD) was analyzed as previously described (Burke et al., 1985Go) to indicate the induction of various cytochrome P450 (Namkung et al., 1988Go).

Histological examination of liver.
Fixed liver tissue was processed into paraffin and 5 µm-thick sections were cut and stained with hematoxylin and eosin. Liver histopathology was analyzed for 5 animals per treatment group by an observer who was unaware of treatment. Sections were scored using an arbitrary scale from 1 to 16 for degree of severity of histopathological lesions. The scale was broken into 4 categories with 1–4 being minimal, 5–8 mild, 9–12 moderate, and 13–16 marked with 1 point increments within each range for focal, diffuse, multifocal, or throughout.

Micronucleus assay of bone marrow.
To assess the degree to which the mixture caused chromosome damage or interfered with mitotic apparatus of developing red blood cells, the incidence of residual chromosome fragments (micronuclei) in polychromatic erythrocytes from femoral bone marrow was determined, as this is a well-established biomarker of chromosome breakage due to DNA damage (Schmid, 1975Go). At necropsy, bone marrow was washed out of the femoral lumen onto a glass slide in 200 µl of bovine serum, smeared, and allowed to air dry. Slides were fixed in methanol and stained with 0.005% acridine orange in 66 mM phosphate buffer (pH 7.3). Stained slides were examined in a dark room using an Olympus fluorescence microscope with mercury lamp illumination. At least 3000 polychromatic erythrocytes were examined for the presence of micronuclei.

Daily sperm production.
Daily sperm production was estimated according to previously described methods (Blazak et al., 1985Go). Briefly, the testis was decapsulated and parenchyma weighed and homogenized in 20 ml of STA solution (0.9 % NaCl, 0.01% Triton X-100, and 0.025% sodium azide) using an Ultramicro blender (Waring, New Hartford, CT) on low setting for 2 min. Testis homogenates were then disrupted with a VibraCell sonicating dismembranator (Sonics & Materials, Danbury, CT). Sperm nuclei density in the testis homogenate was estimated by counting at least 10 replicate samples using a haemocytometer. The number of homogenization resistant testis nuclei was divided by 6.1 days to estimate the number of mature spermatozoa released to the epididydimis per day (Blazak et al., 1985Go). Results were also expressed as the ratio of DSP to testis weight to estimate the efficiency of sperm production (DSP efficiency).

Sperm chromatin structure assay.
The degree of condensation of sperm nuclei is related to fertilizing potential (Evenson and Melamed, 1983Go) and is a parameter indicative of toxicant-induced disruption of spermatogenesis (Evenson, 1986Go; Evenson et al., 1980Go). Sperm chromatin condensation was measured by differential acridine orange fluorescence of acidified sperm, in which condensed and decondensed DNA fluoresce in the green or red ranges respectively. The susceptibility of sperm nuclei to denaturation by acid is estimated by the ratio of red fluorescence to total fluorescence ({alpha}t). The values of {alpha}t for normal sperm fall within a tight range and toxicant-induced damage increases the number of cells falling outside this range such that the percent of cells with {alpha}t more than 2 x SD outside this range (% cells outside of main population) or the variability of {alpha}t in this variant population are inversely correlated with function (Evenson, 1986Go; Evenson et al., 1980Go).

To assess {alpha}t, epididymal sperm suspensions were prepared by homogenizing one cauda epididymis from each animal in ice cold TNE buffer (10 mM Tris, 0.9% NaCl, and 0.5 mM EDTA, pH 7.4) using an Ultra-Turrax homogenizer (Tekmar, Cincinnati, OH) and stored at 4°C for no more than 5 days prior to analysis of sperm chromatin structure by flow cytometry. The Sperm Chromatin Structure Assay (SCSA) utilizes the differential fluorescence of the nucleic acid-binding dye acridine orange (AO) that fluoresces red or green when bound to either denatured or condensed chromatin, respectively. To assess the effect of treatment on sperm chromatin condensation, sperm chromatin structure was assessed in acid-treated sperm nuclei using a flow cytometric method described previously (Foster et al., 1996Go).

Flow cytometric analysis of testis cells.
One fresh testis from each animal was decapsulated into 3 ml of fresh M199 and minced with iris scissors. After all testes were minced, the resulting cell suspensions were filtered through 105 µm diameter plastic mesh. Filtered suspensions were pelleted at 3000 x g for 10 min, resuspended in 70% ethanol, and stored at –20°C for subsequent determination of spermatic cell populations based on DNA staining (Suter et al., 1997Go). Briefly, fixed cells were washed twice in phosphate buffered saline (PBS) and resuspended in PBS. The cell suspension was then passed sequentially through a 25-gauge needle and then a nylon mesh (75 µm mesh opening, Costar Netwell, Costar, Cambridge, MA). A 1 ml aliquot of the filtered cell suspension was stained by adding Tween 20 (0.3% final concentration), RNAase (bovine pancreatic, 10 U/ml, Sigma), and propidium iodide (50 µg/ml, Molecular Probes, Portland, OR) and incubated at 4°C in the dark overnight. Samples were analyzed flow cytometrically with excitation at 488 nm and emmission fluorescence at 630 nm. A minimum of 10,000 cells were examined at a flow rate of 200 cells/s and the results analyzed using List-View (Phoenix Flow Systems, San Diego, CA). Reproducibility of sample manipulation and instrument performance was routinely evaluated for both flow cytometric methods utilized in this study. An aliquot from a single control sample was analyzed for each measurement at least 4 times during each session and coefficients of intra- or intersession variability did not exceed 15% for any measure.

Statistical analyses.
Animals were killed between 0900 and 1300 h on either of 2 consecutive days with the number of animals per treatment group balanced between days. Data that were expressed as percent were normalized by arcsine transformation prior to analysis by ANOVA. Unless otherwise indicated, all data were analyzed by two-way ANOVA with mixture dosage and day of necropsy being the 2 factors tested. Analysis of effects of mixture on variability of {alpha}t was performed using O'Brien's test of homogeneity of variance (O'Brien, 1979Go) Where significant effects were identified for treatment, multiple comparisons were tested using Student Neuman Keuls post hoc test. All data sets were tested for homogeneity of variance and normal distribution and, where homoscedasticity or normality were not indicated, data were retested after log transformation. Where either homoscedasticity or normality tests were not satisfied after this transformation, data were analyzed, for mixture dose treatment effects, but not effect of day of sacrifice, using Kruskal-Wallis ANOVA on ranks test followed by Dunn's test for multiple comparisons. The accepted level of significance was set at p <= 0.05 while tests with a p value between 0.05 and 0.1 were viewed as identifying a trend. All statistical analyses were performed using SigmaStat software version 2.0 (SPSS, Chicago).


    RESULTS
 TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
 DISCUSSION
 REFERENCES
 
Exposure to the mixture did not influence either terminal body weight or weight gain throughout the experiment (data not shown) and there were no signs of stress or discomfort, as indicated by sluggishness, ruffled fur, or lacrimation, in any of the exposed animals. At necropsy, liver and kidney weights were significantly increased by the highest dose of the mixture while there were no effects on the weights of the thymus or adrenal glands, or on most of the organs of the reproductive tract (Table 2Go). There was no indication of genetic toxicity of the mixture as there was no difference in the proportion of bone marrow polychromatic erythrocytes that contained micronuclei as assessed by staining with acridine orange (Table 3Go).


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TABLE 2 Body and Relative Organ Weights (% Body Weight) at Necropsy for Male Rats Exposed to a Complex Mixture of Persistent Environmental Contaminants for 70 Days
 

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TABLE 3 Effects of Mixture Exposure on Incidence of Micronuclei in Bone Marrow Cells
 
To examine general systemic toxicity of the mixture several standard serum biochemical indicators of organ function were assessed (Table 4Go). Of the indicators of liver function assessed, LDH activity was reduced in a dose related fashion, and cholesterol was significantly increased at the highest dose, while a numerical reduction in serum glucose levels was not statistically significant (p = 0.06). In addition, total protein and albumin were increased significantly at the 1000x level. Urea nitrogen was significantly decreased at the 2 highest doses while serum uric acid was significantly reduced only at the highest dose. Serum phosphorus showed a biphasic response with increasing dose of the mixture, being significantly reduced, relative to control levels, in the 10x animals and significantly increased in the animals in the 1000x group. While some of the above indicate mixture-induced changes in liver metabolism, circulating liver enzymes, alkaline phosphatase, and aspartine/leucine transaminase, which could indicate hepatic cell damage, were unaffected by any treatment.


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TABLE 4 Serum Biochemistry at Necropsy for Male Rats Exposed to Mixture
 
To further examine effects on liver drug metabolizing ability, and as a measure of exposure, the activities of several inducible hepatic cytochromes P450 were assessed (Fig. 1Go). EROD and BROD, but not PROD, activities were induced in a dose related fashion by exposure to the mixture. Of these, EROD was the most sensitive showing significant induction of activity in animals receiving the 10x mixture.



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FIG. 1. Effects of 70-day treatment of male rats with a complex mixture of environmental contaminants on EROD (A), PROD (B), and BROD (C) activities measured in S-9 fractions of liver homogenates. Enzyme activities were assessed as described in the text and data are expressed as means and SD of 9 (control) or 10 (all other treatments) animals per group. Means with the same letter superscript are not significantly different (p > 0.05).

 
As relative liver size was increased by exposure to the mixture, histopathological examination of the liver was performed. A variety of histopathological lesions were identified in exposed animals and the incidence and severity of these appeared to increase with increasing dose. In sections from exposed animals, cells tended to be hypertrophied, with less intense eosin staining, such that sinusoids were less evident in sections from highly exposed animals relative to control (Fig. 2Go). In exposed animals hepatocytes in the transition zone between artery and portal vein exhibited increasing degree of vacuolation, suggestive of lipid infiltration, with increased dose. The lesion severity scores appeared to increase with dose with scores (mean ± SD) of 4.00 ± 3.08, 7.60 ± 2.61, 11.20 ± 2.17, 11.25 ± 0.96, and 13.40 ± 3.21 for control, 1, 10, 100, and 1000x treatments, respectively. There was a significant effect of treatment on the lesion severity score data, based on analysis by Kruskal-Wallis ANOVA on ranks, and Dunn's post hoc test demonstrated that only liver samples from the 1000x treatment group showed significantly higher lesion scores compared to control animals.



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FIG. 2. Photomicrograph showing liver from a representative animal administered vehicle (control, panel A) or 1000x contaminant mixture (panel B). Control liver demonstrates normal histology with sinusoid(s) clearly visible between plates of uniform hepatocytes. (B) Liver from high dose animals exhibits pale, hypertrophied hepatocytes adjacent to the portal vessels (P) with many, highly vacuolated cells (arrow heads) in the intermediate region. Bars = 100 µm.

 
Immune Function Effects
To assess immunotoxicity of the mixture, 2 tests of immune function were applied including mitogen-activation of splenic lymphocytes and the determination of NK activity. Spleens collected on the first day of necropsies were not processed until over 42 h after collection due to delays in delivery, which resulted in both quantitative, and in some cases qualitative, differences in response for immune parameters between samples collected on day 1 versus day 2. For those parameters where two-way ANOVA demonstrated significant interaction between treatment effects and day of collection, only data from samples collected on day 2, which were processed within 24 h after collection, are reported here. For Con A, known to stimulate more mature T cells, clonal expansion in response to either 2.5 or 5 µg/ml was highly variable between animals, with coefficients of variation of some treatment means in excess of 100%, and showed large differences between day of collection. However, there was no interaction between factors so data from both days are reported. Con A-induced proliferation was significantly reduced in animals exposed to 1000x mixture relative to all other treatments (Table 5Go). For PHA, known to stimulate less mature T cells, the proliferation in response to 2.5 and 10 µg/ml did show interaction between sampling days. Analysis of data from animals sacrificed on day 2 indicates that the mixture had no significant effect on proliferation stimulated by either 2.5 or 10 µg/ml PHA (Table 5Go).


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TABLE 5 Ex Vivo Functional Responses of Splenocytes Collected from Male Rats Exposed to Contaminant Mixture
 
As two-way ANOVA of NK activity data expressed as cell lysis also showed significant interaction between sampling days, only data from day 2 has been considered. At a dilution of 1/30, treatment with the mixture had a biphasic effect as the lowest dose showed a significant increase in lytic activity, and this response declined with increasing dose (Table 5Go). NK activity in animals from the highest dose was significantly lower than control. At a dilution of 1/60, however, there were no significant differences in NK activity between dose groups (Table 5Go).

Reproductive and Endocrine Effects
There was remarkably little indication of toxicity of the mixture to the wide variety of reproduction-related endpoints examined in this study. Other than effects seen on epididymis weight and epididymal sperm content, no treatment-related effects were seen on the weights of other organs of the male reproductive tract (Table 2Go), daily sperm production (either per testis or per g of testis; Table 6Go), relative numbers of spermatogenic cell populations (Figs. 3A–3DGo), various parameters related to nuclear condensation in sperm (Figs. 4A–4DGo), or serum levels of LH, FSH, or prolactin (Table 7Go). However, exposure to the mixture caused a dose related reduction in both absolute (not shown) and relative total and caput, but not cauda, epididymis weight (Table 2Go).


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TABLE 6 Effects of Subchronic Ingestion of a Complex Mixture of Contaminants on Sperm Production
 


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FIG. 3. The effects of subchronic exposure to a complex of contaminants on the proportions of testicular germ cells that are round spermatids (A), diploid cells (B), cells in S-phase (C), and primary spermatocytes (D). Relative germ cell numbers were based on the intensity of propidium iodide staining of fixed, RNAase-treated testis cells as assessed by flow cytometry. Means with the same letter superscript are not significantly different (p > 0.05).

 


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FIG. 4. The effects of contaminant mixture exposure on chromatin condensation of sperm. Sperm nuclei, prepared from epididymal homogenates by sonication, were stained with acridine orange in an acidic buffer. (A) Alpha t ({alpha}t) is derived as the ratio of red (denatured DNA) to green (condensed double stranded DNA) plus red fluorescence for 10,000 sperm cells per animal and is a measure of sperm maturation or integrity. (B) The SD of {alpha}t is a measure of the average variability in {alpha}t of the 10,000 {alpha}t observations for each animal. Data are expressed as the mean ± SD for all 9 (control) or 10 (all other treatments) animals per group. Means with the same letter superscript are not significantly different (p > 0.05).

 

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TABLE 7 Effects of Mixture Exposure on Levels of Serum Hormones at Sacrifice
 
Flow cytometric analysis of dispersed testicular cells stained for DNA typically yields 4 main peaks: condensed haploid (representing elongated spermatids), round spermatids (round and some elongating spermatids), diploid (G1-phase spermatogonia, preleptotene spermatocytes, and secondary spermatocytes) and primary spermatocytes (contains a minor number of G2/M spermatogonia). In addition to these, cells in S-phase (preleptotene spermatocytes and spermatogonia) can also be resolved between the diploid and primary spermatocyte peaks. In the present study, debris generated through mechanical dispersion of the tissue prevented accurate count of elongate spermatids and, as this population is quantified in measures of daily sperm production, no attempt was made to estimate the relative proportion of this population using flow cytometry. As such the percentages of cells for each population are based only on the total number of cells within the 4 populations that could be routinely resolved. The relative numbers of cells in the round spermatid group were increased in animals treated with the 1000x dose while no effects were seen in this population at other doses (Fig. 3AGo). There was a trend (p = 0.07) towards a reduction in diploid cell number in these high dose animals (Fig. 3BGo) while no differences were seen in any other treatment group in this population or for either s-phase or primary spermatocyte populations (Figs. 3C and 3DGo).

Exposure to the complex mixture had little impact on epididymal sperm chromatin condensation as there were no effects on mean {alpha}t (Fig. 4AGo) or on the variability of {alpha}t values within each animal (Fig. 4BGo). The number of cells for which {alpha}t values differed greatly from the mean value for each animal (percent cells outside the main population; Fig. 4CGo) was significantly reduced in animals in the 10x treatment group, relative to all other groups. In addition these animals showed a slight increase in the SD of this group of cells although this effect was only significant relative to animals from higher dose groups.


    DISCUSSION
 TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
 DISCUSSION
 REFERENCES
 
The current study provides evidence that exposure to low or moderate doses of a mixture of persistent organochlorines and metals affects liver physiology and has some effects on the function of T cells but apparently has little impact on reproductive physiology and does not induce any chromosomal breakage in the adult male rat. As the doses of material administered represent either levels that have been shown to be without effect on the most sensitive endpoint of toxicity known in mammals (no observable adverse effect level or NOAEL) or at levels of subchronic daily exposure well below the NOAEL (i.e., at RfD, TDI, or MRL), the absence of effects seen in the present study at the lowest dose suggests that toxicity of the material to the systems examined is not greatly increased by coexposure to the other mixture components. While the mixture administered does not represent the full panoply of chemical agents to which humans are exposed and, therefore, cannot predict the entire range of chemical interactions leading to toxicity to the endpoint examined, the lack of observed toxicity on these systems at the lowest levels used in this study, in spite of known toxicity of the substances to these systems, offers some support for the assumption that derivation of safe levels of exposure from single agent toxicity studies provide reasonable protection from adverse effects in adult male animals.

The quantities used in formulating the doses of most components of the mixture are promulgated by public environmental health agencies in Canada (Health Canada) or the U.S. (Agency for Toxic Substances and Disease Registry, U.S. Environmental Protection Agency) as guidelines to acquaint physicians or public health professionals with the maximum exposure levels that would not be expected to cause increased risk to health. The sole exception was the dose selected for TCDD that is based on the highest dose of TCDD shown to be without effect on the endpoint shown to be most sensitive to TCDD toxicity (NOAEL of 1 ng/kg/day; Feeley and Grant, 1993Go; Murray et al., 1979Go). Other than TCDD, the safe levels used to formulate the dosing mixtures were calculated using a modification of basic risk assessment methodology used to determine reference dose for lifetime exposure by the U.S. EPA (Barnes and Dourson, 1988Go). This process involves the identification of the highest NOAEL for the most sensitive endpoint in the most sensitive species, where toxicity data are available for multiple species, and dividing this value by an appropriate uncertainty factor. For example, the TDI estimate for 1,2,4-trichlorobenzene, of 0.0023 mg/kg/day, is based on a NOAEL of 22.9 mg/kg/day inhalation exposure for 13 weeks on a variety of general toxicity endpoints (CEPA, 1993Go). This value was then divided by 10 for uncertainty in extrapolating from rodent to human, 10 to account for human variability, 10 for extrapolation from a relatively short-term study to longer-term exposures, and 10 because this study provides the only data on the mammalian toxicity of this compound. The use of these values as a predictor of risk assumes that toxicity will not be modified by coexposure to other substances or that any increase in potency from such interaction will not exceed the safety factor used. For all compounds included in the mixture, with the exception of cadmium, the safe levels are derived from animal studies and have safety factors equivalent to or in excess of 100. The chronic MRL for cadmium is derived from human data on renal toxicity (NOAEL = 2.1 µg/kg/day) corrected by a safety factor of 10 (ATSDR, 1999Go). If one assumes that the available toxicity data for a given compound is a complete reflection of the compound's hazard, the calculated MRL should be less than 1/100 the dose that causes toxicity in the animal models (generally rodents) used to generate the data. In the present study, hepatic EROD activity was significantly elevated in 10x exposed animals. This suggests that the low doses used to derive the mixture formulation were not entirely without effect and implies that the assumption that coexposures are irrelevant is not valid. Alternatively, the endpoints used in determining the NOAEL on which these MRLs are based are less sensitive than the endpoints that were altered by the 10x mixture. It should also be noted that the MRLs used for some of the components have since been replaced with reduced estimates of safe exposure. For example, when the dosing mixture was first formulated, the safe level of dioxin used was based on the interim TDI for PCBs that was 1 µg Arochlor 1254 per kg body weight per day (Grant, 1983Go). Since that time, ATSDR has promulgated a MRL for chronic exposure for PCBs (as Arochlor 1254) of 20 ng/kg/day (ATSDR, 1997Go). In addition, the dose of dioxin of 1 ng/kg/day used in the present study, based not on a MRL type estimate but on a NOEL from a developmental toxicity study, is 3 orders of magnitude above the more recently promulgated MRL of 1.0 pg TEQ/kg/day (ATSDR, 1998Go).

The compounds used to formulate the mixture were chosen to reflect some of the persistent anthropogenic compounds, which have been shown to be present in the tissues of Canadians, to determine if exposure to relatively low levels of mixtures of these may have health consequences. While the lowest doses were below those shown to cause any toxicity, the administered doses are much higher than the estimated average daily exposure to some of these substances for Canadian populations (Birmingham et al., 1989Go; CEPA, 1993Go; Health Canada, 1998Go).

The mixture was clearly hepatotoxic at the highest dose administered as indicated by the increased liver weight and modified liver histology. In addition, hepatic activity of EROD and BROD, but not PROD, were increased in a dose-related fashion. The induction of hepatic EROD activity is characteristic of exposure to arylhydrocarbon receptor agonists such as dioxins or coplanar PCBs (Van den Berg et al., 1998Go), both of which are present in the dosing mixture. The lack of induction of PROD activity suggests that it was not as sensitive as BROD for detecting CYP2B induction. BROD was selected as an endpoint to provide additional indication of cytochrome P450 induction, although the interpretation of the increased catabolism of this substrate due to mixture treatment is difficult as this substrate is also catabolized by CYP1A1, CYP3A, and CYP2B (Namkung et al., 1988Go). As the activity of EROD, exclusively a substrate of CYP1A1 activity, was markedly increased by the mixture, the concomitant increase in BROD activity may simply reflect an increase in CYP1A1 activity, likely in response to the TCDD and, to a lesser extent, the Aroclor in the mixture. The cumulative dose of TCDD to which 10x MRL animals have been exposed (700 ng/kg) is comparable to doses of dioxin that have been shown to induce EROD activity in rats in chronic exposure studies (200 ng/kg, Walker et al., 1999Go) but more than 10-fold higher than the lowest dose shown to induce toxicity to the reproductive tract in rats exposed in utero to a single dose of TCDD on day 15 of gestation (64 ng/kg; Mably et al., 1992aGo,bGo).

Other biochemical parameters sensitive to dioxin or PCB exposure include serum glucose, total protein, and albumin (Chu et al., 1994Go, 2001Go). Depletion of serum glucose is thought to occur via direct inhibition of phosphoenol pyruvate carboxykinase, a key enzyme of gluconeogenesis, by these compounds that ultimately leads to a wasting syndrome (Viluksela et al., 1995Go). The reduction in serum glucose seen in the 1000x animals, although not quite statistically significant, is consistent with many previous studies of relatively long-term dioxin or PCB exposure (e.g., Chu et al., 1994Go; Gorski et al., 1990Go) and will likely impact on intermediary metabolism in these animals. The increase in total protein and albumin may be the result of hypertrophic changes in the liver in response to the mixture and the proliferation of endoplasmic reticulum implied by the obvious cellular hypertrophy (Fig. 2Go) and increased cytochrome P450 activity. Alternatively, increased serum protein and albumin levels could indicate mild dehydration that is consistent with negative water balance seen in rats treated with a single dose of TCDD, similar to the total dose administered to the high dose animals over the 70-day treatment period (50 µg/kg vs. 70 µg/kg in the present study; Potter et al., 1986Go). Other changes seen in the serum are not clearly defined. Increases in calcium and phosphorus and the reduction in uric acid and urea nitrogen may indicate some impaired kidney clearance. This is consistent with the observed increase in kidney weight in the 1000x animals although, in the absence of kidney histopathology or assessment of urinary parameters no definitive conclusions can be drawn. Finally, the reduced serum LDH in the highest doses is interesting and may indicate the inactivation of this enzyme by highly induced mixed function oxidase reactions (Fucci et al., 1983Go) or a direct inhibition by one or more of the mixture components.

In contrast to the toxic effects on liver, the mixture had little effect on the reproductive physiology of adult male animals. This conclusion is based on a broad assessment of reproductive physiology including an examination of several endpoints related to spermatogenesis (relative proportions of testicular cell populations, daily sperm production, epididymal sperm reserves, and sperm nuclear condensation), indicators of endocrine control of reproduction (serum levels of LH, FSH, and prolactin) and terminal weights of reproductive organs. The only organ that appeared to be influenced by the mixture was the epididymis, where the relative weight of the whole or caput were significantly reduced at the two highest doses. In addition, the mixture had an apparently biphasic effect on sperm transit through the epididymis as cauda sperm reserves were significantly increased at the 1x dose but returned to control levels at higher doses. The significance of this nondose-related finding is unclear. Although no statistically significant effects were indicated on the majority of other reproductive endpoints, it should be noted that the power of the statistical tests for most of these endpoints was low. This could imply that endpoints for which the mixture-induced changes were close to (e.g., testis weight) or just at statistical significance (epididymis weights and sperm content) may have shown a stronger response if more animals had been used in the study. Some of the components of the mixture (dioxin, PCB, methoxychlor) have previously been shown to compromise reproductive competence when administered to male rats, although these effects occurred at much higher doses than were present in the dosing mixtures administered in the current study and/or involved the initiation of dosing at a younger age (Chapin et al., 1997Go; Cooke et al., 1996Go; Gray et al., 1989Go, 1995Go; Welshons et al., 1999Go). It should be noted that fetal, neonatal, and immature animals are more sensitive to some mechanisms of action of reproductive toxicants, so the possibility that the mixture may have effects in developing animals is currently being determined in our laboratory.

In addition to reproductive endpoints, the current study also evaluated the mutagenic and immunotoxic effects of the mixture. The effects of the mixture on immune function were also minimal having clear toxic effects only at the highest dose tested. As described above, the ability of the mitogenic assay and NK activity to predict an immunotoxic outcome is 67 and 69% respectively when these tests are used alone. However, when they are used in combination, the predictability increases up to 79% (Luster et al., 1992Go). Despite this, it is very hard to translate immunotoxic outcomes into a direct measurable decrease of resistance to infections; they represent definitely adverse effects representing a serious hazard.

While considerable research has been devoted to the immunotoxicity of single exposures to numerous chemicals, less emphasis has been given to the toxicity of mixtures. Most human exposures to chemicals involve mixtures at relatively low levels, and concurrent exposure to multiple chemicals can increase or decrease the toxicity of single chemicals (Teuschler and Hertzberg, 1995Go). Human volunteers with TCDD body burdens of > 60 ppt, and those with reference TCDD of < 20 ppt, or workers exposed to dibenzofurans showed no evidence of immunosuppression from immunological tests performed (Webb et al., 1989Go; Zober et al., 1992Go). In a previous study, Omara et al. (1997) demonstrated that in vitro exposure of rat lymphocytes to low levels of Aroclor PCB mixtures, PCDD/PCDF mixtures, or MeHg/PCB/PCDD/PCDF mixtures, produced no additive immunotoxicity as assessed by leukocyte cytolethality, and T- and B-cell mitogen-stimulated lymphocyte proliferation, macrophage phagocytic activity, natural killer (NK) and cellular immune response by the one-way mixed leukocyte reaction (MLR). On the other hand other types of mixtures such as particulate air pollution, mixtures of contaminant as present in fish flesh or marine mammals blubber were demonstrated immunotoxic in rodents, especially in mice. The dose range at which certain immunological effects are observed in experimental animals are exposed to chemicals should be taken into consideration before extrapolating data to human exposure situations. A number of observations also suggest that rat and human immune systems appear to be less susceptible than mice to halogenated aromatic hydrocarbons (HAHs), particularly TCDD, which is considered the most toxic of the HAHs (Lang et al., 1994Go; Mocarelli et al., 1986Go; Smialowicz et al., 1994Go; Zober et al., 1992Go).

The results of this study demonstrate that while estimates of safe levels of exposure, applied to a mixture of persistent contaminants, provide apparently good protection for the function of the male reproductive tract, these studies raise concern over their applicability for hepatic and thyroid effects. However, MRL promulgation is a reactive process and the actual values tend to creep downward as new toxicological evidence comes forward as demonstrated by the reduced estimates of safe levels of exposure for PCBs and dioxins. The current study suggests that the MRLs, TDIs, or RfD set by the U.S. ATSDR, U.S. EPA, and CEPA provide adequate protection for adult male animals, for those systems examined, from the adverse effects of these persistent contaminants.


    ACKNOWLEDGMENTS
 
This study was supported in part by a grant from the Canadian Chemical Producers association through the Canadian Chlorine Coordinating Committee and by the Great Lakes Research Program of Health Canada. The authors gratefully acknowledge the technical assistance of Lorraine Casavant and Tarun Ahuja.


    NOTES
 
1 To whom correspondence should be addressed at EHC-315, PL#0803D, Growth and Development Section, Environmental Health Directorate, Tunney's Pasture, Ottawa, Ontario, Canada K1A 0L2. Fax: (613) 946-2600. E-mail: mike_wade{at}hc-sc.gc.ca. Back


    REFERENCES
 TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
 DISCUSSION
 REFERENCES
 
ATSDR (1997). Toxicological Profile for Polychlorinated Biphenyls (Update). Agency for Toxic Substances and Disease Registry, Division of Toxicology, Atlanta, GA.

ATSDR (1998). Toxicological Profile for Chlorinated Dibenzo-p-Dioxins (Update). Agency for Toxic Substances and Disease Registry, Division of Toxicology, Atlanta, GA.

ATSDR (1999). Toxicological Profile for Cadmium (Update). Agency for Toxic Substances and Disease Registry, Division of Toxicology, Atlanta, GA.

Barnes, D. G., and Dourson, M. (1988). Reference dose (RfD): Description and use in health risk assessments. Regul. Toxicol. Pharmacol. 8, 471–486.[ISI][Medline]

Baukloh, V., Bohnet, H. G., Trapp, M., Heeschen, W., Feichtinger, W., and Kemeter, P. (1985). Biocides in human follicular fluid. Ann. N.Y. Acad. Sci. 442, 240–250.[ISI][Medline]

Birmingham, B., Gilman, A., Grant, D. L., Salminen, J., Boddington, M., Thorpe, B., Wile, I., Toft, P., and Armstrong, V. (1989). PCDD/PCDF multimedia exposure analysis for the Canadian population: Detailed exposure estimation. Chemosphere 19, 637–642.[ISI]

Blazak, W. F., Ernst, T. L., and Stewart, B. E. (1985). Potential indicators of reproductive toxicity: Testicular sperm production and epididymal sperm number, transit time, and motility in Fischer 344 rats. Fundam. Appl. Toxicol. 5, 1097–1103.[ISI][Medline]

Boyd, C. A., Weiler, M. H., and Porter, W. P. (1990). Behavioral and neurochemical changes associated with chronic exposure to low-level concentration of pesticide mixtures. J. Toxicol. Environ. Health 30, 209–221.[ISI][Medline]

Brousseau, P., Payette, Y., Blakley, B., Boermans, H., Flipo, D., Tryphonas, H., Fournier, M. (1998). Manual of Immunological Methods. CRC Press, Boston.

Brouwer, A., Morse, D. C., Lans, M. C., Schuur, A. G., Murk, A. J., Klasson-Wehler, E., Bergman, A., and Visser, T. J. (1998). Interactions of persistent environmental organohalogens with the thyroid hormone system: Mechanisms and possible consequences for animal and human health. Toxicol. Ind. Health 14, 59–84.[ISI][Medline]

Burke, M. D., Thompson, S., Elcombe, C. R., Halpert, J., Haaparanta, T., and Mayer, R. T. (1985). Ethoxy-, pentoxy-, and benzyloxy-phenoxazones and homologs: A series of substrates to distinguish between different induced cytochromes p-450. Biochem. Pharmacol. 34, 3337–3345.[ISI][Medline]

CEPA (1993). Trichlorobenzenes. Priority Substances List Assessment Report, Cat. No. En40–215/25E. Canadian Environmental Protection Act, Ministry of Supply and Services Canada, Ottawa.

Chang, L., Gusewitch, G. A., Chritton, D. B. W., Folz, J. C., Lebeck, L. K., and Nehlsen-Cannarella, S. L. (1993). Rapid flow cytometric assay for the assessment of natural killer cell activity. J. Immunol. Methods 166, 45–54.[ISI][Medline]

Chapin, R. E., Harris, M. W., Davis, B. J., Ward, S. M., Wilson, R. E., Mauney, M. A., Lockhart, A. C., Smialowicz, R. J., Moser, V. C., Burka, L. T., and Collins, B. J. (1997). The effects of perinatal/juvenile methoxychlor exposure on adult rat nervous, immune, and reproductive system function. Fundam. Appl. Toxicol. 40, 138–157.[ISI][Medline]

Chapin, R. E., Phelps, J. L., Schwetz, B. A., and Yang, R. S. (1989). Toxicology studies of a chemical mixture of 25 groundwater contaminants. III. Male reproduction study in B6C3F1 mice. Fundam. Appl. Toxicol. 13, 388–398.[ISI][Medline]

Chu, I., LeCavalier, P., Häkansson, H., Yagminas, A., Valli, V. E., Poon, R., and Feeley, M. (2001). Mixture effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin and polychlorinated biphenyl congeners in rats. Chemosphere 43, 807–814.[ISI][Medline]

Chu, I., Villeneuve D. C., Yagminas, A., LeCavalier, P., Poon, R., Feeley, M., Kennedy, S. W., Seegal, R.F., Häkansson, H., Ahlborg, U. G., and Valli, V. E. (1994). Subchronic toxicity of 3,3`,4,4`,5-pentachlorobiphenyl in the rat. I. Clinical, biochemical, hematological, and histopathological changes. Fundam. Appl. Toxicol. 22, 457–468.[ISI][Medline]

Committee on Hormonally Active Agents in the Environment (1999). Hormonally Active Agents in the Environment. National Academy Press, Washington, DC.

Cooke, P. S., Zhao, Y. D., and Hansen, L. G. (1996). Neonatal polychlorinated biphenyl treatment increases adult testis size and sperm production in the rat. Toxicol. Appl. Pharmacol. 136, 112–117.[ISI][Medline]

Davies, D., and Mes, J. (1987). Comparison of the residue levels of some organochlorine compounds in breast milk of the general and indigenous Canadian populations. Bull. Environ. Contam. Toxicol. 39, 743–749.[ISI][Medline]

Evenson, D. P. (1986). Flow cytometry of acridine orange stained sperm is a rapid and practical method for monitoring occupational exposure to genotoxicants. Prog. Clin. Biol. Res. 207, 121–132.[Medline]

Evenson, D. P., Darzynkiewicz, Z., and Melamed, M. R. (1980). Relation of mammalian sperm chromatin heterogeneity to fertility. Science 210, 1131–1133.[ISI][Medline]

Evenson, D. P., and Melamed, M. R. (1983). Rapid analysis of normal and abnormal cell types in human semen and testis biopsies by flow cytometry. J. Histochem. Cytochem. 31(Suppl. A l), 248–253.[ISI][Medline]

Feeley, M. M., and Grant, D. L. (1993). Approach to risk assessment of PCDDs and PCDFs in Canada. Regul. Toxicol. Pharmacol. 18, 428–437.[ISI][Medline]

Foster, W., Chan, S., Platt, L., and Hughes, C. (2000). Detection of endocrine disrupting chemicals in samples of second trimester human amniotic fluid. J. Clin. Endocrinol. Metab. 85, 2954–2957.[Abstract/Free Full Text]

Foster, W. G. (1996). The reproductive toxicity of Great Lakes contaminants. Environ. Health Perspect. 103(Suppl. 9), 63–69.

Foster, W. G., McMahon, A., and Rice, D. C. (1996). Sperm chromatin structure is altered in cynomolgus monkeys with environmentally relevant blood lead levels. Toxicol. Ind. Health 12, 723–735.[ISI][Medline]

Fournier, M., Dégas, V., Colborn, T., Omara, F. O., Denizeau, F., Potworowski, E. F., and Brousseau, P. (2000). Immunosuppression in mice fed on diets containing beluga whale blubber from the St. Lawrence estuary and the Arctic populations. Toxicol. Lett. 112–113, 311–317.[ISI]

Fucci, L., Oliver, C. N., Coon, M. J., and Stadtman, E. R. (1983). Inactivation of key metabolic enzymes by mixed-function oxidation reactions: Possible implication in protein turnover and ageing. Proc. Natl. Acad. Sci. U.S.A. 80, 1521–1525.[Abstract]

Germolec, D. R., Yang, R. S., Ackermann, M. F., Rosenthal, G. J., Boorman, G. A., Blair, P., and Luster, M. I. (1989). Toxicology studies of a chemical mixture of 25 groundwater contaminants. II. Immunosuppression in B6C3F1 mice. Fundam. Appl. Toxicol. 13, 377–387.[ISI][Medline]

Gorski, J. R., Weber, L. W. D., and Rozman, K. (1990). Reduced gluconeogenesis in 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD)-treated rats. Arch. Toxicol. 64, 66–71.[ISI][Medline]

Grant, D. L. (1983). Regulation of PCBs in Canada. In PCBs: Human and Environmental Hazards (F. M. D'Itri and M. A. Kamrin, Eds.), pp. 383–392. Butterworth, Boston.

Gray, L. E., Jr., Kelce, W. R., Monosson, E., Ostby, J. S., and Birnbaum, L. S. (1995). Exposure to TCDD during development permanently alters reproductive function in male Long Evans rats and hamsters: Reduced ejaculated and epididymal sperm numbers and sex accessory gland weights in offspring with normal androgenic status. Toxicol. Appl. Pharmacol. 131, 108–118.[ISI][Medline]

Gray, L. E., Jr., Ostby, J. S., Ferrell, J., Rehnberg, G. L., Linder, R., Cooper, R. L., Goldman, J. M., Slott, V., and Laskey, J. (1989). A dose-response analysis of methoxychlor-induced alterations of reproductive development and function in the rat. Fundam. Appl. Toxicol. 12, 92–108.[ISI][Medline]

Gyorkos, J., Denomme, M. A., Leece, B., Homonko, K., Valli, V. E., and Safe, S. H. (1985). Reconstituted halogenated hydrocarbon pesticide and pollutant mixtures found in human tissues: Effects on the immature male Wistar rat after short-term exposure. Can. J. Physiol. Pharmacol. 63, 36–43.[ISI][Medline]

Health Canada (1998). Persistent Environmental Contaminants and the Great Lakes Basin population: An exposure assessment. Cat. No. H46–2/98–218E. Government of Canada, Ottawa.

Heindel, J. J., Chapin, R. E., George, J., Gulati, D. K., Fail, P. A., Barnes, L. H., and Yang, R. S. (1995). Assessment of the reproductive toxicity of a complex mixture of 25 groundwater contaminants in mice and rats. Fundam. Appl. Toxicol. 25, 9–19.[ISI][Medline]

Heindel, J. J., Chapin, R. E., Gulati, D. K., George, J. D., Price, C. J., Marr, M. C., Myers, C. B., Barnes, L. H., Fail, P. A., Grizzle, T. B., et al. (1994). Assessment of the reproductive and developmental toxicity of pesticide/fertilizer mixtures based on confirmed pesticide contamination in California and Iowa groundwater. Fundam. Appl. Toxicol. 22, 605–621.[ISI][Medline]

Ito, N., Hagiwara, A., Tamano, S., Hasegawa, R., Imaida, K., Hirose, M., and Shirai, T. (1995). Lack of carcinogenicity of pesticide mixtures administered in the diet at acceptable daily intake (ADI) dose levels in rats. Toxicol. Lett. 82–83, 513–520.

Jarrell, J. F., Villeneuve, D., Franklin, C., Bartlett, S., Wrixon, W., Kohut, J., and Zouves, C. G. (1993). Contamination of human ovarian follicular fluid and serum by chlorinated organic compounds in three Canadian cities. CMAJ 148, 1321–1327.[Abstract]

Lang, D. S., Becker, S., Clark, G. C., Devlin, R. B. and Koren, H. S. (1994). Lack of immunosuppressive effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) on human peripheral blood lymphocyte subsets in vitro. Arch. Toxicol. 68, 296–302.[ISI][Medline]

Lapierre, P., De Guise, S., Muir, D. C. G., Norstrom, R., Beland, P., and Fournier, M. (1999). Immune functions in the Fisher rat fed beluga whale (Delphinapterus leucas) blubber from the contaminated St. Lawrence estuary. Environ. Res. 80, S104–S112.[ISI][Medline]

Liu, J., Liu, Y., Barter, R. A., and Klaassen, C. D. (1995). Alteration of thyroid homeostasis by UDP-glucuronosyltransferase inducers in rats: A dose-response study. J. Pharmacol. Exp. Ther. 273, 977–985.[Abstract]

Lubet, R. A., Nims, R. W., Mayer, R. T. Cameron, J. W., and Schechtman, L. M. (1985). Measurement of cytochrome P-450 dependent dealkylation of alkoxyphenoxazones in hepatic S9s and hepatocyte homogenates: Effects of dicumarol. Mutat. Res. 142, 127–131.[ISI][Medline]

Luster, M. I., Portier, C., Pait, D. G., White, K. L., Jr., Gennings, C., Munson, A. E., Rosenthal, G. J., (1992). Risk assessment in immunotoxicology. I. Sensitivity and predictability of immune tests. Fundam. Appl. Toxicol. 18, 200–210.[ISI][Medline]

Mably, T. A., Bjerke, D. L., Moore, R. W., Gendron-Fitzpatrick, A., and Peterson, R. E. (1992a). In utero and lactational exposure of male rats to 2,3,7,8-tetrachlorodibenzo-p-dioxin. 3. Effects on spermatogenesis and reproductive capability. Toxicol. Appl. Pharmacol. 114, 118–126.[ISI][Medline]

Mably, T. A., Moore, R. W., and Peterson, R. E. (1992b). In utero and lactational exposure of male rats to 2,3,7,8,-tetrachlorodibenzo-p-dioxin. 1. Effects on androgenic status. Toxicol. Appl. Pharmacol. 114, 97–107.[ISI][Medline]

Mes, J. (1992). Organochlorine residues in human blood and biopsy fat and their relationship. Bull. Environ. Contam. Toxicol. 48, 815–820.[ISI][Medline]

Mes, J., and Malcolm, S. (1992). Comparison of chlorinated hydrocarbon residues in human populations from the Great Lakes and other regions in Canada. Chemosphere 25, 417–424.[ISI]

Mes, J., and Marchand, L. (1987). Comparison of some specific polychlorinated biphenyl isomers in human and monkey milk. Bull. Environ. Contam. Toxicol. 39, 736–742.[ISI][Medline]

Mes, J., Marchand, L., and Davies, D. J. (1990). Organochlorine residues in adipose tissue of Canadians. Bull. Environ. Contam. Toxicol. 45, 681–688.[ISI][Medline]

Mocarelli, P., Marocchi, A., Brambilla, P., Gerthoux, P., Young, D. S., and Mantel, N. (1986). Clinical laboratory manifestations of exposure to dioxin in children. A six-year study of the effects of an environmental disaster near Seveso, Italy. JAMA 256, 2687–2695.[Abstract]

Murray, F. J., Smith, F. A., Nitschke, K. D., Humiston, C. G., Kociba, R. J., and Schwetz, B. A. (1979). Three-generation reproduction study of rats given 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in the diet. Toxicol. Appl. Pharmacol. 50, 241–252.[ISI][Medline]

Namkung, M. J., Yang, H. L., Hulla, J. E., and Juchau, M. R.(1988). On the substrate specificity of cytochrome P450IIIA1. Molec. Pharmacol. 34, 628–637.[Abstract]

O'Brien, R. G. (1979). A general ANOVA method for robust tests of additive models for variances. J. Amer. Stat. Assoc. 74, 877–880.[ISI]

Omara, F. O., Brochu, C., Flipo, D., Denizeau, F., and Fournier, M. (1997). Immunotoxicity of environmentally relevant mixtures of polychlorinated aromatic hydrocarbons with methyl mercury on rat lymphocytes in vitro. Environ. Toxicol. Chem. 16, 576–581.[ISI]

Omara, F. O., Flipo, D., Brochu, C., Denizeau, F., Brousseau, P., Potworowski, E. F., and Fournier, M. (1998). Lack of suppressive effects of mixtures containing low levels of methylmercury (MeHg), polychlorinated dibenzo-p-dioxins (PCDDs), polychlorinated dibenzofurans (PCDFs), and aroclor biphenyls (PCBs) on mixed lymphocyte reaction, phagocytic, and natural killer cell activities of rat lymphocytes in vitro. J. Toxicol. Environ. Health A 54, 561–577.[Medline]

Omara, F. O., Vincent, R., and Fournier, M. (2000). The inhibitory effect of urban air pollution particles on mitogen-activated rat and mouse lymphocyte proliferation is reversible by N-acetylcysteine. J. Toxicol. Env. Health 59, 101–119.

Pauwels, A., Covaci, A., Delbeke, L., Punjabi, U., and Schepens, P. J. (1999). The relation between levels of selected PCB congeners in human serum and follicular fluid. Chemosphere 39, 2433–2441.[ISI][Medline]

Potter, C. L. Menahan, L. A., and Peterson, R. E. (1986). Relationship of alterations in energy metabolism to hypophagia in rats treated with 2,3,7,8-tetrachlorodibenzo-p-dioxin. Fundam. Appl. Toxicol. 6, 89–97.[ISI][Medline]

Schlebusch, H., Wagner, U., van der Ven, H., al Hasani, S., Diedrich, K., and Krebs, D. (1989). Polychlorinated biphenyls: The occurrence of the main congeners in follicular and sperm fluids. J. Clin. Chem. Clin. Biochem. 27, 663–667.[ISI][Medline]

Schmid, W. (1975). The micronucleus test. Mutat. Res. 31, 9–15.[ISI][Medline]

Singh, S. K., and Pandey, R. S. (1989). Gonadal toxicity of short term chronic endosulfan exposure to male rats. Indian J. Exp. Biol. 27, 341–346.[ISI][Medline]

Smialowicz, R. J., Riddle, M. M., Williams, W. C., and Diliberto, J. J. (1994). Effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) on humoral immunity and lymphocyte subpopulations: Differences between mice and rats. Toxicol. Appl. Pharmacol. 124, 248–256.[ISI][Medline]

Suter, L., Koch, E., Bechter, R., and Bobadilla, M. (1997). Three-parameter flow cytometric analysis of rat spermatogenesis. Cytometry 27, 161–168.[ISI][Medline]

Teuschler, L. K., and Hertzberg, R. C. (1995). Current and future risk assessment guidelines, policy, and methods development for chemical mixtures. Toxicology 105, 137–144.[ISI][Medline]

Tryphonas, H., Fournier, M., Lacroix, F., McGuire, P., Hayward, S., Bryce, F., Flipo, D., and Arnold, D. L. (1998a). Effects of Great Lakes fish consumption on the immune system of Sprague-Dawley rats investigated during a two-generation reproductive study. Part II: Quantitative and functional aspects. Regul. Toxicol. Pharmacol. 27, 40–54.

Tryphonas, H., McGuire, P., Fernie, S., Miller, D., Stapley, R., Bryce, F., Arnold, D. L., and Fournier, M. (1998b). Effects of Great Lakes fish consumption on the immune system of Sprague-Dawley rats investigated during a two-generation reproductive study. Part I. Body and organ weights, food consumption and haematological parameters. Regul. Toxicol. Pharmacol. 27, S28–S39.[ISI][Medline]

Van den Berg, M., Birnbaum, L., Bosveld, A. T., Brunström, B., Cook, P., Feeley, M., Giesy, J. P, Hanberg, A., Hasegawa, R., Kennedy, S. W., Kubiak, T., Larsen, J. C., van Leeuwen, F. X., Liem. A. K., Nolt, C., Peterson, R. E., Poellinger, L., Safe, S., Schrenk, D., Tillitt, D., Tysklind, M., Younes, M., Waern, F., and Zacherewski, T. (1998). Toxic equivalency factors (TEFs) for PCBs, PCDDs, PCDFs for human and wildlife. Environ. Health Persp. 106, 775–792.[ISI][Medline]

Van Hove Holdrinet, M., Braun, H. E., Frank, R., Stopps, G. J., Smout, M. S., and McWade, J. W. (1977). Organochlorine residues in human adipose tissue and milk from Ontario residents 1969–1974. Can. J. Publ. Health 68, 74–80.[ISI]

Viluksela, M., Stahl, B. U., and Rozman, K. K. (1995). Tissue-specific effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) on the action of phosphoenolpyruvate carboxykinase (PEPCK) in rats. Toxicol. Appl. Pharmacol. 135, 308–315.[ISI][Medline]

Walker, N. J., Portier, C. J., Lax, S. F., Crofts, F. G., Li, Y., Lucier, G. W., and Sutter, T. R. (1999). Characterization of the dose response of CYP1B1, CYP1A1, and CYP1A2 in the liver of female Sprague-Dawley rats following chronic exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol. Appl. Pharmacol. 154, 279–286.[ISI][Medline]

Webb, K. B., Evans, R. G., Knutsen, A. P., Roodman, S. T., Roberts, D. W., Schramm, W. F., Gibson, B. B., Andrews, J.S., Jr., Needham, L. L., and Patterson, D. G. (1989). Medical evaluation of subjects with known body levels of 2,3,7,8-tetrachorodibenzo-p-dioxin. J. Toxicol. Environ. Health 28, 183–193.[ISI][Medline]

Welshons, W. V., Nagel, S. C., Thayer, K. A., Judy, B. M., and Vom Saal, F. S. (1999). Low-dose bioactivity of xenoestrogens in animals: Fetal exposure to low doses of methoxychlor and other xenoestrogens increases adult prostate size in mice. Toxicol. Ind. Health 15, 12–25.[ISI][Medline]

Zober, M. A., Ott, M. G., Papke, O., Senft, K., and Germann, C. (1992). Morbidity study of extruder personnel with potential exposure to brominated dioxins and furans. I. Results of blood monitoring and immunological tests. Br. J. Ind. Med. 49, 532–544.[ISI][Medline]