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a California Regional Primate Research Center, University of California at Davis, Davis, California 95616-8542
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ABSTRACT |
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 TOP ABSTRACT INTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONREFERENCES |
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INTRODUCTION |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONREFERENCES |
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Previous studies from our group have shown that the endocrine signals of early pregnancy in the cynomolgus macaque (Macaca fascicularis) and human are quite similar [3]. The macaque embryo implants about 9 days after fertilization, while the human embryo implants 3 days earlier. Trophoblast-ovarian signaling appears to be similar in the two species. The appearance of chorionic gonadotropin (CG) occurs approximately 3 days after implantation in both species and coincides with the acceleration of secretion of RLX. Increased estrogen production by the ovary, which is also a response to CG, is comparable, i.e., Day 9 in the human and Day 12 in the monkey. The only significant difference in these early pregnancy signals is the time course of CG production by the trophoblast, which begins to decline by 20 days in macaques, while in humans it continues to increase [4].
The present study was performed to document the endocrine changes that occur when EFL is induced in primates after exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). There have been several well-documented human exposures to dioxin-like chemicals, including polychlorinated biphenyl (PCB) exposures in Japan and Taiwan [5], and dioxin exposures in Seveso, Italy [6]. Exposure to dioxins results in a number of toxic and biological responses [7], and TCDD is often studied as a representative chemical for many dioxin-like compounds. Moreover, preliminary data are available on TCDD toxicity in the rhesus macaque, a species closely related to the cynomolgus macaque. McNulty [8] reported a high incidence of abortions in rhesus macaques when TCDD was administered at oral doses of 1.0 µg/kg during early pregnancy, although maternal toxicity was delayed.
The present study was undertaken to collect experimental data on the endocrine biomarkers that may be most useful for detecting toxicant-induced EFL in human populations. One difficulty in associating EFL with environmental exposures is the relatively high rate of spontaneous pregnancy failure in nonexposed women [1]. Approximately one-third of all conceptions end in spontaneous abortion, with two-thirds of these being EFL, which is detected only by the appearance of immunoreactive hCG. Only an increase in pregnancy loss over the 30% spontaneous rate would be indicative of toxicant-induced losses since the pattern of hCG production in environmentally induced losses and spontaneous losses are not known to be different. A biomarker that identified abnormalities of pregnancy immediately after exposure to an environmental hazard would improve our capability to detect environmentally induced pregnancy losses in prospective epidemiologic studies.
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MATERIALS AND METHODS |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONREFERENCES |
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Female cynomolgus monkeys ranging in age from 6 to 12 yr old and weighing 3.1 to 4.3 kg, with a known reproductive history, were used in this study. All animals were housed and maintained according to current guidelines for laboratory animal care. Animal rooms, maintained on a 12L:12D cycle (lights-on at 0600 h), and a year-round temperature of approximately 22ÅC and 60% humidity, were equipped with an automatic watering system for provision of water ad libitum. Animals were fed Purina Monkey Chow (Ralston-Purina Co., St. Louis, MO) twice daily. The California Regional Primate Research Center (CRPRC) is fully accredited by the American Association for the Accreditation of Laboratory Animal Care (AAALAC), and all research protocols and procedures related to these studies were approved by the University of California, Davis Animal Use and Care Committee, the Campus Veterinarian, and the CRPRC.
Ovulation Detection and Pregnancy Monitoring
Daily early morning urine collections were initiated on cycle Day 7 (cycle Day 1 was the first day of vaginal bleeding) through Day 15 to monitor for the urinary estrone conjugates (E1C) peak (which is an indicator of ovulation). The day of the E1C peak was designated as gestational day (GD) 0. Blood sample collections were initiated on GD 7 and continued every day until GD 28, then every third day until GD 40. Animals were bred on cycle Days 10 and 12, and this breeding interval was verified to encompass the periovulatory E1C peak. Pregnancy was confirmed on GD 12 by the measurement of > 1 ng/ml serum CG [9] and by trans-abdominal ultrasound [10] (Ultramark 9; Advanced Technology Laboratories, Inc., Bothell, WA). There were no false positives or negatives using CG measurements and sonography [10]. Nonpregnant animals were rescheduled for another breeding at this time. Estradiol (E2), progesterone (P4), and RLX were measured in all serum samples. In the case of pregnancy loss, the final blood sample was collected on the day the loss was detected by sonography. In cases of EFL, some embryos were removed by hysterotomy between GDs 27 and 30 for gross and histopathological examination. Follow-up ultrasound examinations in pregnant animals were performed daily through GD 20 and then on GDs 25, 30, 35, and 60 and approximately every 3 wk thereafter for fetuses that survived exposure [11, 12].
Hormone Assays
Serum E2 and P4 concentrations were determined by methods previously described [13, 14]. Circulating RLX and immunoreactive CG concentrations were determined by heterologous and homologous ELISAs as previously described ([3] and [9], respectively). Urinary metabolites of estrogen, E1C, were determined as previously described by this laboratory [15]. Circulating bioactive CG concentrations were measured by the in vitro bioassay described by Jia et al. [16] with modifications as described below. The interassay coefficients of variation for the serum E2, P4, RLX, immunoreactive CG, bioactive CG, and urinary E1C assays in this study were 15%, 14%, 12.2%, 10%, 14%, and 15%, respectively.
In Vitro Bioassay for CG
For the CG bioassay, clonal human fetal kidney cells (cell line 293) were cultured in Dulbecco's Modified Eagle Medium (DMEM; Gibco BRL, Grand Island, NY) containing 5% calf serum, 2 mM L-glutamine, 100 U/ml penicillin, 100 µg/ml streptomycin sulfate, and 100 µg/ml geneticin sulfate. The cells were cotransfected with pGL2-inh
-luc, pCMXhLHR, and a plasmid (pSV2neo) containing the gene for neomycin resistance. After transfection, the cells were cultured in media containing G418 (Gibco BRL, Gaithersburg, MD). Several single-cell colonies from transfected cells were isolated, and their luciferase activity and LH receptor content were measured. One cell colony (DL293) had both receptor binding and luciferase activity and was used for development of the LH/CG bioassay. After the DL293 cells were cultured to 80–100% confluence in 100 x 20-mm tissue culture dishes (Applied Scientific, South San Francisco, CA), cells were counted and incubated in Micro-well cell plates (Nunclon TM Surface; Applied Scientific). Each well contained 1–10 x 105 cells in 100 µl DMEM supplemented with basic fibroblast growth factor (Intergen Company, Purchase, NY) and 0.05 mM 3-methyl isobutyl/xanthine (MIX; Sigma Chemical Co., St. Louis, MO). Then 10 µl hCG CR 127 standards (provided by R. Canfield, Columbia University, NY), internal controls (pooled monkey serum), and experimental serum samples for assay were added. LH/CG bioactivity in each sample was measured in duplicate. In order to ensure a constant volume of 10 µl serum in the total volume of 110 µl, all assay samples and hormone standards were balanced with calf serum. At the end of incubation (20 h), 110 µl double-strength lysis buffer (single-strength solution: 25 mM Tris-phosphate pH 7.8, 2 mM dithiothreitol (DTT), 2 mM 1,2-diaminocyclohexane-N,N,N',N'-tetra-acetic acid (cDTA), 10% glycerol, 1% Triton X-100) was added. The plates were incubated at room temperature for 1 h to allow cell lysis. For estimation of luciferase activity, 30 µl of the cell lysate was transferred to Micro Fluor plates (Dynatech Laboratories, Inc., Chantilly, VA), and then mixed with 100 µl assay buffer (0.5 mM luciferin, 20 mM Tricine (Sigma), 1.07 mM [(MgCO3)4Mg(OH)2]·5H2O, 2.67 mM MgSO4, 0.1 mM EDTA, 33.3 mM DTT, 0.27 mM coenzyme A, and 0.5 mM ATP). Light production was measured for 2 sec by a microtiter plate luminometer (Dynatech Laboratories, Inc.). LH/CG bioactivity in serum samples was calculated by use of a standard curve using CR 127. The assay sensitivity was determined by the dose of LH/CG capable of stimulating luciferase activity greater than the mean plus 2 x SE of the basal light production by untreated cultures. The intraassay error, expressed as coefficient of variation of pooled monkey serum was 8.04% (n = 30).
TCDD Treatment
Crystalline TCDD was obtained from Cambridge Isotope Labs (Andover, MA) at 99% purity. TCDD was dissolved in acetone (100 µg/ml), and serial dilutions of this stock solution were made with corn oil to prepare dosing solutions of 1, 2, and 4 µg/ml for the 1-, 2-, and 4-µg/kg dose groups, respectively. The dosing solutions were vortexed for 15 min in sterile vials protected from light with aluminum foil. The dosing volume was 1 ml TCDD solution/kg BW; the controls received the same dosing volume of acetone-corn oil. TCDD or vehicle was administered as a single dose by nasogastric intubation under yellow light conditions. Animals were treated once with TCDD (treatment Day 0) on the day of pregnancy detection (GD 12) [10].
To determine the effects of TCDD on early pregnancy, 19 pregnant animals were dosed; 7 served as controls and received the solvent/vehicle (acetone-corn oil) only. The remaining 12 animals were divided into low- (1 µg/kg), mid- (2 µg/kg), and high- (4 µg/kg) dose treatment groups of 4 animals each (Table 1). The low dose (1 µg/kg) was chosen for reference to previous studies carried out in rhesus monkeys by McNulty [8]. In the initial experiment with 4 animals, 1 µg/kg administered on GD 12 elicited only 2 cases of EFL. Thus, the 2- and 4-µg/kg groups were added to increase the incidence of EFL. These doses are within the range of doses used in previous TCDD studies in various species [17].
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Maternal/Infant Observations
Maternal weights were measured once predosing and once postdosing (within 7 days before and after dosing) and biweekly thereafter. All animals were observed daily for clinical signs. Complete blood counts and clinical chemistry panels were performed on maternal blood samples pretreatment, as well as 1 wk and 1 mo posttreatment. Standard necropsies were performed on all animals that died during the study. The surviving pregnancies were monitored according to established protocols [12]. The surviving infants were weighed, sexed, measured, and examined for external malformations. Comparisons were made between TCDD-exposed infants and concurrent and historical controls for the cynomolgus monkey colony.
Biohazard Procedures
Because of the toxicity of TCDD, stringent safety procedures were followed to prevent hazardous exposure to laboratory personnel. These included the use of protective clothing and gloves during all phases of the study, decontamination of all supplies by autoclaving or exposure to ultraviolet light for 24 h, housing of animals in metabolism cages for complete collection of urine and feces, and posted emergency procedures for accidental exposures. Disposal of TCDD was carried out according to California Environmental Protection Agency regulations, which stipulate limits for nonbiohazardous disposal of liquid and solid TCDD waste at 1 µg/L and 10 µg/kg, respectively.
Statistics
The comparison of all hormone values (immunoreactive CG, bioactive CG, E2, P4, and RLX) and of ratios of immunoreactive:bioactive CG values between treatment groups and the control group were analyzed by repeated-measures ANOVA. For serum E2, P4, and RLX, individual values were averaged across weeks to address the problem of missing values; when one or more values were missing, data that were present were just averaged and then log-transformed before analysis. The log transformations were necessary in order to diminish the impact of outliers in these kinds of data. Basically, the log transformation was used to address the fact that the residual variability increased as the mean increased, and the need to reduce the impact of outliers is reasonable justification for this transformation. Data are presented as mean ± SEM. p < 0.05 was considered to be significant.
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RESULTS |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONREFERENCES |
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Poor appetite was recorded at variable times posttreatment in the majority (9 of 12) of treated females. Eight animals (n = 2 low dose, n = 2 mid dose, and n = 4 high dose) experienced weight loss (13–23% decrease in pretreatment weight) up to 8 wk posttreatment. A variety of posttreatment clinical signs indicative of TCDD-induced maternal toxicity were also observed across all dose groups. These signs first occurred 10–82 days after TCDD administration. These indications of maternal toxicity were generally observed after embryonic death, which occurred 10–20 days posttreatment (GDs 22–32). The discrepancy in detecting onset of maternal and embryonic toxicity may be due to the sensitivity of the methods used for detecting toxicity (i.e., clinical signs versus ultrasound). Alopecia occurred in one animal in each dose group; dermatitis was additionally observed in one mid-dose and one high-dose animal. Various changes in the eye and eyelid, including swelling, excessive tearing, ocular discharge, inflammation, and loss of eyelashes were observed in one low-dose animal, two animals in the mid-dose group, and three animals in the high-dose group. Partially and completely sloughed fingernails occurred in one low-dose animal and two high-dose animals. Posttreatment anemia was commonly observed in the high-dose group (n = 4) and low-dose group (n = 2). No emesis or diarrhea was observed. Three animals died during the study, one in the low-dose group (28 days posttreatment) and one in the mid-dose group (58 days posttreatment). A third animal died approximately 1 yr after exposure to the high dose.
Developmental Toxicity Following TCDD Treatment
The pregnancy outcomes following TCDD exposure in the periimplantation period are shown in Table 1. No EFL occurred in the seven animals in the control group, while embryonic death followed by abortion was observed in 10 of the 12 animals in the TCDD-treated groups. In the 4 animals in the low-dose group, two embryos were sonographically identified as nonviable on GDs 31 and 32. One animal aborted all products of conception 14 days after embryonic death was detected (GD 46), and one animal was found dead 9 days after embryonic death was noted. The outcomes of the pregnancies that did not abort were a premature stillbirth on GD 142, and a normal term infant that was delivered vaginally.
In the mid-dose group, 4 of 4 treatments resulted in embryonic death (GDs 24–32) and subsequent abortion. One animal was noted with a viable embryo on GD 21 and an abortion-in-progress on GD 24. The remaining three animals were observed with normal, viable embryos 2–3 days before detection of embryonic demise (GDs 31–32) (see Table 1), and aborted all products of conception from 7–18 days after embryonic death was identified sonographically (GDs 38–50). The animals in the high-dose group also had early pregnancy termination in 4/4 cases. One animal had a complete abortion on GD 22 (last observation of a viable embryo on GD 19). The remaining three animals were also observed with normal, viable embryos 2–3 days before detection of embryonic demise (GDs 26–32); complete abortions occurred 4–28 days later (GDs 30–60). Thus, assessments of all dose groups indicate that embryonic death (10 of 12) occurred on GDs 22–26 in 3 of 10 (30%), and on GDs 31–32 in 7 of 10 (70%) animals. The gross pathology of several embryos examined shortly after death indicated extensive congestion of the major blood vessels and heart as well as accumulation of blood in the pericardial cavity. Histopathology demonstrated excessive cell death in major regions of the brain, developing vertebrae, and the gut epithelium.
TCDD Effects on Endocrine Biomarkers
Two different patterns of circulating immunoreactive CG were observed depending on the dose of TCDD administered. In the mid- and high-dose groups, immunoreactive CG levels were slightly lower than those of controls, particularly between posttreatment Days 7 and 12, after which the levels were more similar to control values and remained unchanged for the remainder of the observation period (Fig. 1). In contrast, circulating levels of immunoreactive CG were higher than control in the low-dose treatment group; particularly in the two animals with surviving embryos, in which CG levels were approximately twice as high as control values.
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In contrast to the profiles of immunoreactive CG, bioactive CG began to rise a day later in the treated animals compared to controls and reached peak levels significantly lower than that of the controls independent of dose (Fig. 2). By Day 9 posttreatment, the mean bioactive CG concentration was approximately 3-fold higher in control animals than the means of all three treatment groups. In general, bioactive CG remained significantly lower in all treatment groups compared to controls (p < 0.01) until posttreatment Day 19, when levels in both treated and controls were near baseline. There was also a significant difference in the level of bioactive CG between those animals that maintained their pregnancy in the low-dose group and the animals that did abort, independent of treatment group (p < 0.05; Fig. 2).
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When the bioactive to immunoreactive (B:I) CG ratio was calculated for each day for each animal, that ratio was relatively low for the first six days in all groups (Fig. 3). After posttreatment Day 6, the mean B:I ratio of the control group rose rapidly to reach a peak ratio on posttreatment Day 10. In contrast, the B:I ratio did not rise significantly in any treatment group (regardless of pregnancy outcome). The B:I ratio was significantly lower (p < 0.001) in all treated animals than in control animals whether or not an EFL was induced.
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Serum E2 concentrations were highly variable between individual animals but tended to reveal a biphasic rise in all animals with surviving pregnancies. TCDD treatment tended to block both the early rise (Days 3–7 [GDs 15–19]) and the later rise (Day 12 [GD 24]), and concentrations were significantly lower in animals receiving TCDD compared to controls (p < 0.02) (Fig. 4). No significant differences were found with any other hormone, but there was a nonsignificant trend for a reduction in serum P4 (p = 0.10) and serum RLX (p < 0.08).
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DISCUSSION |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONREFERENCES |
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Whether or not the endocrine changes that followed TCDD treatment were the cause of EFL, these biomarkers revealed evidence of reproductive toxicity almost immediately after exposure. Surprisingly, the profiles of immunoreactive CG did not reveal the toxic exposure. The most striking endocrine alteration associated with TCDD exposure was the reduction in bioactivity of CG. In our previous reports describing changes in the B:I ratio of hCG during early human pregnancies, two well-characterized immunoassays for the intact hCG dimer were employed [18, 19]. These methods and the use of the same standards in the immunoassays and bioassay enabled us to make direct comparisons between the amount of immunoreactive versus bioactive hCG in each sample. The immunoassay for monkey CG in the present study does not have a sensitivity or specificity comparable to the immunoassays that are available for hCG. For example, the immunoassay for monkey CG does not detect pregnancy until approximately Day 12 postovulation in macaques [9], whereas hCG assays detect human pregnancy by Day 9 postovulation [4]. These limitations are unavoidable at the present time because monkey CG has not been isolated, purified, or fully characterized, and thus specific antibodies to the molecule have not been generated. For this reason, the higher levels of the bioactive forms of CG in comparison with the immunoreactive forms should be interpreted with caution. However, by comparison to control values using the same assays, it is apparent that TCDD did not suppress the production of immunoreactive CG at any dose (Fig. 1), whereas the effect of TCDD on bioactive CG was profound. Bioactive CG began to rise a day later in the treated animals than in control animals, and by Day 9 posttreatment bioactive CG was 3-fold higher in controls than in treated animals (Fig. 2). Bioactive CG continued to be lower in treated animals than in controls until treatment Day 19, when levels in both treated and control groups were near baseline (Fig. 2). Although there was variation between animals, the profiles of bioactive CG were similar in the different treatment groups and in animals that aborted versus those that maintained their pregnancies (Fig. 2). These data suggest either that TCDD has an all-or-none effect at the level of the trophoblast or that even the lowest dose was sufficient to cause the maximal effect on production of bioactive CG.
When the B:I CG ratios were calculated for control and treated groups, the B:I ratios of both groups were relatively low for the first six days after treatment (Fig. 3). After six days, the B:I ratio of control animals rose rapidly to reach a peak value on treatment Day 10. In contrast, the B:I ratio in treated animals remained low throughout the period of CG production (Fig. 3). In the low-dose group, there was a paradoxical increase in production of immunoreactive CG after TCDD treatment. This response cannot be explained by any known regulatory mechanism since CG production is thought to be autonomous at this time. One possibility is that the lower dose of TCDD may have stimulated CG production by the trophoblast cells either through an effect of TCDD on protein kinase A activity, which has been documented in other tissues [20], or through other mechanisms that changed the rate of trophoblast invasion or differentiation. A failure to observe this effect at higher doses maybe attributed to a dose-dependent increase in cytotoxicity that began to be apparent at the higher doses. The consistent decrease in bioactive CG concentration, however, argues for a specific effect on the structure of the CG molecule as well as an effect on its production. To our knowledge, this kind of endocrine disruption by a polychlorinated aromatic hydrocarbon has not been observed previously and may represent a specific posttranslational target of toxicity in the trophoblast cell.
The significant decrease in E2 production observed in all treatment groups is consistent with a direct effect of TCDD on ovarian function. While this effect on steroidogenesis could be secondary to effects on CG bioactivity, this explanation seems unlikely since there was no similar decrease in RLX production. The increase in the secretion of RLX and E2 during normal early pregnancy is a result of stimulation by CG, acting most likely on cells within the corpus luteum. The disproportionate effect of TCDD in reducing E2 secretion in comparison with P4 and RLX secretion suggests a specificity of toxic action either to different cell types within the corpus luteum or to different biochemical targets within the same cell types.
The data from control animals in this study show that during the normal periimplantation period in macaques, there is a gradual increase in the ratio of bioactive to immunoreactive CG that is coincident with the overall increase in CG production by the implanting trophoblast. In animals that are treated with TCDD and are destined to abort, there is a reduction in bioactive CG, but not immunoreactive CG. These experimental findings are consistent with clinical observations that bioactive hCG, but not immunoreactive hCG, is reduced in the immediate postimplantation period of human pregnancies that later abort spontaneously [18]. The slight increase in immunoreactive CG observed in the present study may be mechanistically related to the observation that treatment of trophoblast cells with two other aryl-hydrocarbon receptor (AHR) ligands (benzo[a]pyrene and 3-methylcholanthrene) resulted in increased hCG production in vitro [21]. The present data also are consistent with the results of experiments in vitro demonstrating that normal growth and development of the human trophoblast is associated with secretion of hCG molecules of increasing bioactivity [19], and that treatment of human trophoblast cells with TCDD in vitro alters both the normal development of the cells and their production of bioactive hCG (unpublished data).
In summary, the results of this study demonstrate the toxic effect of TCDD on early primate pregnancy. Although the multiple effects of TCDD on several endocrine organs could not be associated with a single target of toxicity, the perturbations of the endocrine biomarkers provided striking evidence of toxic exposure many days before there was any other sign of reproductive toxicity. Interestingly, the toxic exposure was not revealed by the profile of immunoreactive CG, which is widely used in epidemiologic studies as the biomarker for prospective assessment of human pregnancy. The present results also demonstrate that normal pregnancy in the monkey, as in the human, is characterized by a postimplantation change in the B:I ratio of CG, and that this profile is altered in pregnancies that are destined to abort. These findings suggest that the macaque is the relevant animal model with which to conduct experiments on the physiologic and biochemical basis of this previously unrecognized aspect of embryo-maternal signaling in primates. This monkey model can also be used to demonstrate the utility of monitoring changes in the production of bioactive CG as a biomarker of environmental exposures that lead to EFL.
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ACKNOWLEDGMENTS |
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FOOTNOTES |
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2 Correspondence: Bill Lasley, California Regional Primate Research Center, University of California, One Shields Avenue, Davis, CA 95616-8542. FAX: 530 752 5300; bllasley{at}ucdavis.edu
Accepted: October 21, 1998.
Received: July 31, 1998.
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONREFERENCES |
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