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Regular Article |
a Department of Urology (M/C 955), College of Medicine, University of Illinois, Chicago, Illinois 60612-7310
b Department of Environmental & Molecular Toxicology, North Carolina State University, Raleigh, North Carolina 27695-7633
c Endocrinology Branch, Reproductive Toxicology Division, National Health and Environment Effects Research Laboratory MD72, United States Environmental Protection Agency, Durham, North Carolina 27713
ABSTRACT
Environmental contaminants with estrogenic properties have been cause for heightened concern about their possible role in inducing adverse health effects. Brief exposure of rodents to high doses of natural estrogens early in life results in permanent alterations of the male reproductive tissues, but the question of whether environmentally relevant doses can cause the same effects remains controversial. The current project was designed to determine the dose-response relationship between neonatal estradiol exposure and the development of the male reproductive tract in the rat. Neonatal male Sprague-Dawley (SD) and Fisher 344 (F344) rats were exposed to ß-estradiol-3-benzoate (EB) at concentrations ranging from 0.015 µg/kg body weight (BW) to 15.0 mg/kg BW and 0.15 µg/kg BW to 1.5 mg/kg BW, respectively. Results showed an inverted U-shaped dose-response profile for testis and epididymis weights in 35-day-old SD rats, with increased organ sizes at the low-dose end of the treatment. This effect was transient and was not sustained into adulthood. Increased hepatic testosterone hydroxylase activities in low-dose animals suggest an advancement of puberty as the cause for increased reproductive organ weights. On postnatal day (PND) 90, a stimulatory low-dose response to EB was present in SD rat testicular and epididymal weights, however at one order of magnitude lower dose than that seen on PND 35, suggesting a separate effect. All SD male reproductive tract organs and serum hormones showed a permanent inhibitory response to high doses of neonatal EB. F344 rats exhibited greater estrogen sensitivity on PND 90. Despite this heightened responsiveness, F344 rats did not exhibit a low-dose effect for any endpoint. These low-dose responses to estradiol are organ and strain specific.
early development, environment, estradiol, prostate, puberty, steroid hormones, toxicology
INTRODUCTION
Recent reports about possible disruption of development and reproductive functions caused by a variety of environmental contaminants with estrogenic or antiandrogenic properties have heightened public concerns about the adverse health consequences of these chemicals for many species, including humans. In wildlife populations, developmental and reproductive derangements such as testicular feminization, alteration of plasma testosterone concentrations, poor survivorship, declining birthrates, reduced phallus sizes, and reduced hatching success have been observed [1–5]. Although controversial [6], exposure to estrogenic chemicals is considered a possible cause of locally increased rates of testicular cancer, declining sperm counts and semen quality, cryptorchidism, and hypospadias in humans [7]. Furthermore, xenoestrogens have been suggested to affect prostate development [8, 9]. In light of this rapidly growing body of evidence, the endocrine disruptor hypothesis has become a major focus of research aimed at improving existing risk assessments.
There is consensus that xenoestrogens can disrupt endocrine functions when present at high concentrations and during specific sensitive periods of development, as was the case for fetal in utero exposure to diethylstilbestrol (DES), a synthetic estrogen that was given to great numbers of pregnant women until the 1970s [10]. However, environmentally relevant levels of hormonally active chemicals usually lie far below those considered to result in adverse effects and may therefore pose little or no threat at all. This traditional dose-response concept of toxicology, with a threshold below which no adverse effect can be observed and with a monotonic response curve, has been challenged both theoretically [11] and experimentally [12, 13]. The alternative no-threshold model holds that there exists no threshold dose if the effect caused by an exogenous chemical is mediated through the same mechanism as that associated with an endogenous chemical. As opposed to increasing toxicity being proportional to increasing concentrations of a toxicant, in a nonmonotonic or U-shaped dose response, low-dose exposure leads to the inverse of the effect caused by high-dose exposure. If confirmed, these two models would require a reevaluation of existing risk assessment protocols. It is therefore of pivotal importance to test both hypotheses in various experimental in vivo systems.
Prenatal and postnatal exposure to low environmentally relevant concentrations of estrogenic chemicals have produced a range of developmental effects such as sex reversal in turtles [13], reduced testes sizes and Sertoli cell numbers in Wistar rats in response to DES [14], enhanced induction of cytochrome P-4501A activity by pentachlorbiphenyl in mice following low doses of tributylin [15], increased anogenital distances in CD-1 mice in response to low doses of DES, bisphenol A, and aroclor 1016 [16], increased preputial gland size caused by bisphenol A in mice [17], and increased reproductive tract organ sizes after in utero exposure to low doses of estradiol or DES [12, 18]. In the first part of the present study, we investigated the effects of estradiol benzoate (EB) on the development of the rat prostate gland and observed a transient increase of relative dorsal prostate weights caused by neonatal exposure to low doses of EB [9]. The transient nature of the effect and its absence in adult animals suggested a possible advancement of puberty. Others have observed precocious puberty of female mice indicated by a reduction of number of days between first vaginal opening and first vaginal estrus in response to low concentrations of bisphenol A administered prenatally [19, 20], which suggests that low doses of estrogens may affect the embryonic imprinting of central processes involved in regulating the onset of sexual maturity.
The goals of the present study were twofold. First, we sought to establish the full dose-response relationship between neonatal EB exposure and the effect on the entire male reproductive tract along with other hormonally relevant tissues in the rat model. Both the Sprague-Dawley (SD) rat, a strain conventionally used in toxicology, and the estrogen-sensitive Fisher 344 (F344) strain were examined to address the issue of previously reported differential strain responses to endocrine-modulating chemicals [21]. Since initial findings suggested that puberty may be advanced in response to low doses of EB in the rat, our second objective was to characterize this event by monitoring reliable pubertal markers. Preputial separation, the androgen-dependent separation of the prepuce from the glans penis (balanus) due to balanopreputial epithelium cornification, occurs at the time of male puberty and has been widely used as a puberty marker [22, 23]. Therefore, we monitored the age at preputial separation and the anogenital distance on the first day of preputial separation in estrogen-exposed rats. Alterations in hepatic testosterone biotransformation enzymes, which are specific for either male or female puberty [24, 25], also were measured in exposed animals of different ages. Serum hormone levels were measured on postnatal days (PNDs) 35 and 90 to determine whether these parameters were differentially affected by low or high doses of estrogen.
MATERIALS AND METHODS
Animals and Housing
All animals were handled in accordance with the principles and procedures of the Guiding Principles for the Care and Use of Animal Research, and their use in the procedures was approved by the Institutional Animal Care and Use Committee at the University of Illinois College of Medicine. Timed-pregnant SD rats (Zivic-Miller Laboratories, Pittsburgh, PA) and F344 rats (Harlan Teklad, Madison, WI) were housed individually in a light- and temperature-controlled environment (ambient temperature approximately 21°C and 60% relative humidity, 14L:10D schedule). All animals had free access to a special soy-free diet (Sterol Free Rat Diet, Ziegler Bros., Gardners, PA) and to tap water at all times. The same lots of rat diet were fed in both experiments (SD and F344) to control for variation in estrogenicity of different rat diet batches.
Females were observed daily and were left undisturbed until one day after parturition (PND 1), when the total numbers of live and dead pups and of male and female pups were recorded for each litter. Litters of SD dams were culled randomly on PND 1 to a maximum of 10 neonates, with as many as 10 male pups per litter if possible. Litters containing fewer than 10 males were completed to 10 with the appropriate numbers of female siblings. In litters with less than 10 pups, nothing was altered. The litters of F344 females were not culled, and all male offspring available were recruited.
In-Life Observations
Because animals were monitored throughout the entire experiment, animals were marked for individual identification on PND 1 using the toe-clipping technique. After weaning on PND 21, up to three male animals at a time were housed in one cage until they were killed on either PND 35 (block 1) or PND 90 (block 2). Individual body weights (BWs) and anogenital distances (AGDs; distance between anus and the base of the phallus) were recorded on PNDs 7, 14, 21, 28, 35, 65, and 90. From PND 28 onwards, animals were checked for preputial separation (PPS). Animals were considered to have reached PPS when the prepuce could be manually retracted entirely over the glans using only gentle pressure.
Experimental Design
The design of the experiments is described in detail elsewhere [9]. In the first experiment, 32 pregnant female SD rats were randomly assigned to a treatment group. Neonates were treated on PNDs 1, 3, and 5 with a 7-log range of doses (0.015 µg kg BW-1 day-1 to 15 mg kg BW-1 day-1) of EB (Sigma Chemical Co., St. Louis, MO) by s.c. injections. In the second experiment, six timed-pregnant F344 dams were randomly assigned to one of five treatment groups (naive control, oil control, and 0.15, 15, or 1500 µg EB kg BW-1 day-1) and were monitored for birth. Male neonates were injected s.c. on PNDs 1, 3, and 5 as described for SD pups. Animals were killed on PNDs 35 and 90 (SD experiment) or on PND 90 (F344 experiment) by cervical dislocation. Testes, epididymes, seminal vesicles, coagulating glands, ampullary glands, kidneys, adrenal glands, and livers were removed from each animal, weighed, snap frozen in liquid nitrogen, and stored at -80°C. All dissections were performed by an operator who was blind to the treatment group.
Serum Hormone Titers
Serum hormone titers were determined by RIA for testosterone, LH, and FSH. Iodination preparations (1-6 for LH and FSH), reference preparation (RP-3), and antisera (S-8 for LH, S-11 for FSH) were provided by the National Institute of Arthritis, Diabetes, Digestive and Kidney Diseases. The purified iodination preparations were radiolabeled with 125I (New England Nuclear, Boston, MA) using the chloramine-T method of Greenwood et al. [26]. Labeled hormones were separated from unreacted iodide on a BioGel P-60 (BioRad Laboratories, Hercules, CA) column with potassium phosphate buffer (50 mM, pH 7.5). For the assay, goat anti-rabbit gamma-globulin (Calbiochem, La Jolla, CA) was used as the second antibody. The assay sensitivities of testosterone, LH, FSH, and the reference preparation were increased by a 24-h coincubation of sample and first antibody prior to the addition of the labeled hormone. The respective sensitivities and inter- and intra-assay coefficients of variation for the two hormones were 0.1 ng/tube, 7.3%, and 6.1% for testosterone and 15 pg/tube, 7.7%, and 5.9% for LH.
Liver Enzyme Activities
Testosterone hydroxylase and oxido/reductase activities were measured as described previously [27]. Assays were performed using 400 µg microsomal protein and 40 nmol [14C]testosterone (New England Nuclear) as substrate (1.8 µCi/µmol) in 0.1 M potassium phosphate buffer (pH 7.4). Reactions were conducted for 10 min at 37°C and initiated with 1 mM NADPH. Resulting hydroxylated derivates of testosterone were extracted with ethyl acetate and separated by thin-layer chromatography. Individual metabolites were quantified using an autoradiography image analyzer.
Statistics
Comparison of mean first day of PPS, AGD, enzyme activity, BW, relative organ weights (mg/100 mg BW), and serum hormone titers for various treatment groups was made using a one-way ANOVA after establishing homogeneity of groups. Treatment groups were compared against the control group in a post hoc Dunnett multiple comparison test. If necessary, a nonparametric Kruskal-Wallis ANOVA was employed to compare medians, followed by the Dunn multiple comparison test. Probability values of <0.05 were accepted as significant. Unless stated otherwise, all values are expressed as mean ± SEM with indication of the number of separate determinations corresponding to individual numbers.
RESULTS
Response of SD Rats to Neonatal EB Treatment
Effects of treatment on BW and reproductive organ weights BWs differed significantly between treatment groups of block 1 and block 2 but not in a dose-dependent manner (Table 1). On PND 35, two groups had lower BWs than did control animals (150.0 and 1500.0 µg/kg BW, P < 0.01), whereas on PND 90 two different groups had reduced BWs (0.015 and 15 000.0 µg/kg BW, P < 0.05). In contrast, reduction of AGD in treated animals on PND 35 was dose dependent and occurred only in high-dose animals treated with 1500.0 or 15 000.0 µg/kg BW on PND 35 (P < 0.05), an effect that in case of the higher dose persisted in 90-day-old animals (P < 0.01).
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A nonmonotonic dose-dependent response was observed in testes on PND 35, where low doses of EB (0.15 µg/kg BW) increased (P < 0.05) and high doses of EB (1500.0 and 15 000.0 µg/kg BW) decreased (P < 0.01) relative organ weights significantly (Fig. 1A). Relative epididymal weights responded monotonically on PND 35, with no significant (P > 0.05) changes at low and medium concentrations of EB but with significantly (P < 0.05) reduced weights at 1500.0 and 15 000.0 µg/kg BW (Fig. 1B). Although this low-dose effect was present on PND 90 (Fig. 1, B and D), it shifted between PND 35 and PND 90 from animals treated with 0.15 µg/kg BW to those treated with 0.015 µg/kg BW. This low-dose response shift suggests that these effects may be independent and manifested through separate pathways on PND 35 and PND 90. In contrast to testis and epididymis, the seminal vesicle and coagulating gland responded in a monotonic dose-dependent fashion, with decreasing organ sizes in response to increasing doses of EB and no alterations in size at low EB concentrations (Fig. 2). Adrenal gland and kidney weights were obtained for all groups, because both organs have been observed to be sensitive to estrogen, androgen, or combined effects (unpublished data) [28]. High doses of EB (1500.0 and 15 000.0 µg/kg BW) significantly increased relative but not absolute adrenal weights on PND 35. This effect was permanent in animals treated with the highest dose of EB (P < 0.01). In contrast, relative kidney weights were unaffected on both PND 35 and PND 90, whereas absolute weights were significantly lowered on PND 90 in animals exposed to 1500.0 and 15 000.0 µg/kg BW (Table 1).
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Effects of treatment on pubertal markers Onset of PPS as an external marker for puberty was only determined for block 2 animals (90-day experiment), because none of the animals of block 1 (35-day experiment) reached PPS prior to being killed. At lower doses, EB had no effect on the development of PPS (P > 0.05; Fig. 3A), whereas in animals given high doses (1500 and 15 000 µg/kg BW) prepuce and balanus did not separate at all prior to PND 90. Body weights at time of PPS were significantly increased in the mid-dose range (1.5 to 150.0 µg/kg BW; P < 0.05; Fig. 3B), although not in a dose-dependent manner. Likewise, AGD at PPS was only significantly increased in one treatment group (1.5 µg/kg BW; P < 0.05; Fig. 3C) and did not show any consistently dose-related alterations. In the entire SD experiment, only one case of cryptorchidism was observed (0.54% of 184 animals) in an animal from one of the low-dose (1.5 µg/kg BW) groups.
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On PND 35, the overall shape of the dose-response curve for several hepatic testosterone biotransformation enzymes was nonmonotonic (Fig. 4). Mid-range doses of EB (1.5 and 15.0 µg/kg BW) increased activities of hepatic microsomal 2
-, 16-, 7-, and 6ß-testosterone hydroxylases (OHTs), whereas high doses reduced enzymic activities (Fig. 4, A–D). The amplitudes of the observed changes were greatest for 2-OHT and were significant (P < 0.05); all other testosterone hydroxylases showed the same trend but not at significant levels (P > 0.05). In contrast, 17ß-hydroxysteroid dehydrogenase (17ß-OHSD) was unaltered (P > 0.05) but 5-reductase activity was increased in animals treated with high concentrations of EB.
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On PND 90, the nonmonotonic dose response of 2- and 16 -OHT shifted at the low end so that only animals treated with the lowest concentration of EB (0.015 µg/kg BW) had increased enzymic activities (Fig. 4, A and B), whereas at the high doses, treatment caused permanent and significant suppression of 2- and 16 -OHT activities in the liver (P <0.05). Similar depression of activity was observed for 6ß-OHT (Fig. 4D), but not for 7-OHT, where no permanent effect was recognizable (Fig. 4C). 17ß-OHSD was reduced in a monotonic dose-dependent fashion with significant decreases in high-dose animals (P < 0.05). The approximately fourfold elevation of 5-reductase activity in animals treated with high concentrations of EB was significant (1353.11 ± 984.09 pmol min-1 mg-1 in controls to 3421.11 ± 480.77 pmol min-1 mg-1 in the 1500.0 µg/kg BW group, P < 0.05, and to 4401.61 ± 2114.99 pmol min-1 mg-1 in the 15 000.0 µg/kg BW group, P < 0.05).
Serum hormone levels Serum LH levels on PND 35 were not significantly different between the treatment groups, although the mean value was increased at the lowest EB concentration and decreased at the highest concentration (Fig. 5A). By PND 90, no significant differences in serum LH titers between groups were observed (Fig. 5B). Mean testosterone serum levels on PND 35 were elevated in the lowest treatment group, although not significantly (Fig. 5C). On PND 90, the dose-response curve for serum testosterone resembled an inverted U-shaped curve, but only the decreased levels at the highest EB concentration were significant (Fig. 5D).
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Response of F344 Rats to Neonatal EB Treatment
Effects of treatment on reproductive organ weights High doses of EB reduced BWs of F344 rats on PND 90 significantly (P < 0.01; Table 2). Accordingly, AGDs in animals of the same treatment group were likewise shortened significantly (P < 0.01). Unlike in SD rats, relative testes and epididymes weights were only affected in a negative manner by the highest concentration of EB (P < 0.05; Fig. 6, A and B). F344 rats were far more sensitive to the estrogen treatment than were SD rats. In the 1500.0 µg/kg BW treatment group, relative testicular weights were reduced to 44.16% of the control weights in contrast to reduction to 67.37% of control weights in SD rats. Likewise, epididymal sizes in F344 rats were reduced to 36.05% of controls as opposed to 86.62% of controls in SD rats. This difference in estrogen sensitivity was even more pronounced in the seminal vesicle and coagulating gland weights (Fig. 6, C and D), where weight reductions were observed at only the highest EB concentration and low doses had no permanent effect. While relative seminal vesicle weights in F344 rats were reduced to only 7.88% of control weights, they were 14.95% of control weights in SD rats. The mid-dose of 15.0 µg/kg BW, which had caused a significant reduction in SD rats, did not alter mean relative seminal vesicle weights in F344 rats, further demonstrating the differential response of different strains to estrogen exposure. Coagulating glands were reduced by an even larger percentage to 10.11% of control weights in F344 rats, as opposed to 31.99% of control weights in SD rats. Ampullary gland weights were not determined for SD rats. However, at the highest treatment dose in F344 rats, the development of the glands was suppressed so greatly that they could not be sampled from 90-day-old animals. At low and mid doses, relative ampullary gland weights were not significantly different from control weights (P > 0.05; Fig. 6E).
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As observed previously for SD rats, relative adrenal gland weights were significantly increased in high-dose animals (P < 0.001), although absolute adrenal gland weights were not (Table 2). The difference between relative control and high-dose adrenal weights was slightly smaller in F344 rats than in SD rats (116.29% of control weights in F344 rats as opposed to 125.86% of control weights in SD rats). Spleen weights did not differ significantly between treatment groups.
Effects of treatment on pubertal markers Similar to SD rats, high-dose F344 rats (1500.0 µg/kg BW) did not reach PPS prior to the termination of the experiment on PND 90 (Table 3). Neither low nor mid doses of EB affected age, BWs, or AGD at PPS compared with control animals (P > 0.05).
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DISCUSSION
Perinatal and neonatal exposures to various concentrations of natural and synthetic estrogens cause irreversible organizational changes in the developing male rodent reproductive tract. In particular, high doses of estrogens have been demonstrated to cause such profound permanent effects as feminization, reduction of organ weights, reduced sperm production, morphologic alterations of the prostate gland, and infertility [29–31]. In the present study, low doses of natural estrogens caused the inverted manifestation of some of these effects, thus resulting in nonmonotonic dose-response profiles, which in some cases were present only transiently in the younger animals but in others remained into adulthood. The temporary character of some of these changes and the transient increase of hepatic testosterone hydroxylases suggest that low doses of EB led to precocious puberty, whereas high doses resulted in the known estrogenization effects associated with organizational imprinting [32]. Furthermore, direct comparison of SD and F344 rats confirmed previously reported strain differences in response to estrogen exposure. Thus, despite the greater estrogen sensitivity of the F344 rat, certain low-dose effects observed transiently in the less sensitive SD rat were not sustained in the adults, confirming the temporal nature of these effects.
The weight data for testes and epididymides revealed nonmontonic dose-response profiles and imply two independent consequences of neonatal low-dose estrogen treatment. On PND 35, relative testicular and epididymal weights were increased at 0.15 µg/kg BW (although not significantly in the epididymis), an effect that, like relative prostate weights in the same animals [9], was transient and did not persist until PND 90. However, at 0.015 µg/kg BW weights of both organs were significantly increased in the adult animals, revealing a permanent effect not observed previously for the prostate gland. This shift by one logarithmic treatment dose suggests two possibly separate effects, of which the transient one may be best explained by an advancement of puberty.
Puberty can be defined as the transition into adulthood, when mature gametes are first produced and reproductive activity is initiated [33]. This transitional process is associated with increased gonadal sex steroid production in response to augmented secretion of gonadotropins from the anterior pituitary gland. High concentrations of exogenous natural or synthetic estrogens administered during early postnatal life can advance the pubertal process in females [19, 34] but delay its onset in males [29]. Low-dose estrogen, however, has been reported to induce early puberty in males; ethinyl estradiol significantly accelerated the puberty marker PPS at concentrations of 5 and 25 ppb in rodent diet [35]. In the present study, animals that received high doses of EB (1500.0 or 15 000.0 µg/kg) early in life had not developed PPS by adulthood, suggesting feminization. However, the dose-response curve showed a threshold between 150.0 and 1500.0 µg/kg, below which EB had no apparent effect on the onset of PPS. Nonetheless, the age at PPS was slightly increased by concentrations of 1.5 µg/kg or higher, the same doses that caused increased BWs at PPS. Since PPS and BW are positively correlated [36], this finding indicates that PPS occurred with a slight delay in animals neonatally exposed to 1.5–150.0 µg/kg EB. Both serum LH and testosterone titers were increased in animals treated with the lower and middle doses, suggesting that although not manifested as PPS puberty may have occurred precociously in response to the treatment. Despite its wide use as a puberty marker, PPS is only a general indicator of the exact onset of the pubescent phase [23], so that the observed small differences in age at PPS may be less definitive than the other, more sensitive physiologic indicators.
Far more precise indication for the onset of puberty than PPS is provided by measuring hepatic testosterone metabolizing enzyme activities, which are also altered in response to neonatal steroid exposure [37]. Various steroid hydroxylating and reducing enzymes show sexually dimorphic expression in the rat liver such as 5-reductase, which has a higher activity in the female, and CYP2C11 (2- and 16-hydroxylase), which is expressed at higher levels in the male [25, 38]. Previous exposure studies have clearly demonstrated that the evolution of CYP2C11 expression and the metabolism of its substrate in liver occurs in response to an early neonatal imprinting of the hypothalamus by testosterone [38, 39]. Once normally imprinted, the hepatic enzymic activity of CYP2C11 increases rapidly by approximately 10-fold during puberty in the male, whereas in the female its expression remains low [40]. Sonderfan et al. [24] showed that the selective expression of cytochrome P-450 in the liver largely accounted for this pubertal increase in 2- and 16-testosterone oxidation in males. In rats as in most mammals, growth hormone (GH) release becomes sexually dimorphic during puberty, with males displaying high amplitude, low frequency GH pulsations and females exhibiting low amplitude, high frequency GH pulses with elevated interpulse baselines [41]. This male pattern of GH secretion has been identified as the mechanism that induces CYP2C11 [42].
We observed a demasculinization in the hepatic enzyme activity patterns of prepubertal males treated with high concentrations of EB, whereas low-dose animals were supermasculinized, suggesting advanced puberty. Others have reported demasculinization of the male hepatic steroid metabolism pattern in response to high doses of estrogens [37], but to our knowledge this is the first observation that neonatal exposure to low doses of estrogen can stimulate 2- and 16-hydroxylase activities in the pubescent male rat liver. The differential pulsatile GH secretion pattern in males and females is programmed neonatally by the action of estradiol at the hypothalamus [43], which may explain why high-dose males were feminized, as indicated by the significant increase of 5-reductase activity and the decrease in 2- and 16-hydroxylase activities. However, supermasculinization of 35-day-old male rats that had been neonatally exposed to mid-range concentrations of EB indicates that neonatal imprinting occurred in a fashion similar to that of testosterone imprinting, thus acting indirectly on hepatic enzyme expression via the hypothalamus, which controls GH secretion in the rat [42]. If that hypothesis were correct, GH spurts of pubertal amplitude would occur in low-dose animals of prepubescent age (i.e., PND 35) and CYP2C11 would be induced, resulting in higher enzyme levels in the liver. The transient nature of the observed increased enzyme activities in livers of mid-dose rats lends further support to the hypothesis of precocious puberty, although the dose range at which this stimulation occurred did not correspond directly with the dose range that resulted in increased relative organ weights. This difference may reflect the gradual advancement of puberty as a function of the treatment dose.
In the male pubescent rat, increasing testosterone levels precede the 10-fold elevation of GH and its pulsatility during puberty [43], thus causing 2- and 16-hydroxylase activities to rise by a margin of similar dimensions [40]. Prostate, testes, and epididymides weights also increase during puberty, although their reaction to higher androgen levels may not be as immediate as that of GH and, in turn, hepatic 2- and 16-hydroxylase. Hence, in the present study, increased enzymic activity and increased organ weights would not necessarily occur simultaneously in the same animals (Fig. 7). On PND 35, hepatic 2- and 16-hydroxylase activities in lower dose animals (0.15 µg/kg) may have already decreased to adult values, whereas the response on the organ level was evident. However, animals of the next higher treatment group (1.5 µg/kg) displayed the first signs of puberty (i.e., increased enzyme activities) with a slight delay in comparison to the lower dose group yet still precocious in comparison to control animals. At that point, their organ sizes would not have been increased and would therefore still be at control levels. The asymmetric occurrence of increased hepatic testosterone hydroxylase activities and reproductive organ weights thus reflects a gradual temporal and dose-dependent onset of puberty.
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It has long been known that administration of estrogen to male rats affects reproductive organs differently, depending on the age of the animal, dosage, and duration of the treatment [44]. Early experiments showed distinct differences in responses, particularly of the prostate, seminal vesicle, and coagulating gland, that were characterized not only by the degree of induced effects but also by which tissues reacted [45]. Hence, disparate responses observed in the present study, such as the monotonic, dose-dependent decrease in size of seminal vesicles and coagulating glands as opposed to nonmonotonic responses of prostate [9], testes, and epididymides or the sustained low-dose effects on the testis and epididymis, may reflect differential sensitivity of these organs to their hormonal environment. A possible explanation for such organ-specific discrepancies in response to estrogen exposure could be the embryologic origin of the organs in question. The epididymis, vas deferens, and seminal vesicle have Wolffian duct origins [45], whereas both the prostate and the coagulating gland are derived from epithelial outgrowths of the urogenital sinus [46]. The specific mesenchyme into which the epithelia of both the Wolffian duct and urogenital sinus grow are essential for the subsequent differentiation of the resulting organs, making the epithelium-mesenchyme interaction crucial for organogenesis [47]. This interaction is orchestrated by several developmental genes that are regulated by steroids [48, 49]. Neonatal low-dose estrogenization may alter their expression along the developing Wolffian duct and urogenital sinus, which in turn could be responsible for the observed organ-specific responses.
Differential sensitivity of SD and F344 rats to neonatal estrogen treatment in the present experiment supports recent observations of genetic predisposition for estrogen responsiveness in different rodent strains [50]. In a direct comparison of the responses of the two rat strains to estrogenic chemicals, female F344 rats had been reported to be more sensitive than SD rats to estrogen treatment [51–53]. However, not only do the two strains differ in their responsiveness to estrogen exposure, but they generally display significant variation in their physiology, development, behavior, and age-related changes in physiologic parameters [54–58]. Strain differences in age-related changes have been reported to be organ specific. Inano et al. [59] reported a greater incidence of mammary tumors in aging male SD than F344 rats implanted with DES pellets at 2 months of age. In light of these studies, the differential sensitivity of male F344 rats to neonatal estrogenization observed in the present investigation must be cautiously considered with regard to gender and the observed endpoints. It may be inaccurate to declare the F344 rat the more reliable and more sensitive sentinel strain for the assessment of xenoestrogens; yet, the current results at least suggest that for investigations into the response of the developing reproductive tract of male rats to neonatal estrogen exposure, F344 rats and SD rats have markedly different reactions.
The present work bears particular importance for the current endocrine disruptor controversy and its pressing question of whether low-dose effects and nonmonotonic dose-response curves really exist. It is essential to identify endpoints that may respond to lower doses of estrogens than do others, thus making them physiologic markers for potentially adverse xenoestrogen effects. Our results suggest that puberty, and especially the onset of the pubescent period, may be such an important marker for low-dose estrogen effects in males. Since previously used puberty markers such as PPS appear less reliable, we propose hepatic testosterone biotransformation enzymes as more accurate indicators for puberty, which therefore should be included in risk assessment protocols for probable xenoestrogens. Reports that various organochlorine pesticides were able to induce testosterone hydroxylases in the liver of the rat further support this conclusion [60]. The rat strain used for risk assessment of estrogenic chemicals should ideally be highly sensitive to moderate or preferably very small concentrations, as was the case for F344 rats when compared with SD rats in the present study. However, because the two strains show a great variability in their biology even under normal conditions, the choice of a model for assessing the effects of endocrine-disrupting chemicals will depend on the gender and specific endpoints.
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ACKNOWLEDGMENTS
We thank Lynn Birch for her excellent technical assistance and Selena Mistich and Keith McElroy for their excellent assistance with the RIAs.
FOOTNOTES
1 Supported by EPA STAR grant R826299 to G.S.P. 
2 Correspondence: Gail S. Prins, Department of Urology (M/C 955), University of Illinois, 820 South Wood Street, Chicago, IL 60612-7310. FAX: 312 996 1291; gprins{at}uic.edu
Accepted: July 3, 2001.
Received: March 27, 2001.
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 H.O. Goyal, A. Robateau, T.D. Braden, C.S. Williams, K.K. Srivastava, and K. Ali Neonatal Estrogen Exposure of Male Rats Alters Reproductive Functions at Adulthood Biol Reprod, June 1, 2003; 68(6): 2081 - 2091. [Abstract] [Full Text] [PDF]
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R. Thuillier, Y. Wang, and M. Culty Prenatal Exposure to Estrogenic Compounds Alters the Expression Pattern of Platelet-Derived Growth Factor Receptors {alpha} and {beta} in Neonatal Rat Testis: Identification of Gonocytes as Targets of Estrogen Exposure Biol Reprod, March 1, 2003; 68(3): 867 - 880. [Abstract] [Full Text] [PDF]
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J. J. Bianco, D. J. Handelsman, J. S. Pedersen, and G. P. Risbridger Direct Response of the Murine Prostate Gland and Seminal Vesicles to Estradiol Endocrinology, December 1, 2002; 143(12): 4922 - 4933. [Abstract] [Full Text] [PDF]
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