Reproductive toxicity evaluation of dietary butyl benzyl phthalate (BBP) in rats

Reproductive Toxicology 18 (2004) 241–264

Reproductive toxicity evaluation of dietary butyl benzyl phthalate (BBP) in rats Rochelle W. Tyl a,∗ , Christina B. Myers a , Melissa C. Marr a , Patricia A. Fail a , John C. Seely b , Dolores R. Brine a , Robert A. Barter c , John H. Butala d a

RTI International, 245 HLB/MCB, P.O. Box 12194, 3040 Cornwallis Road, Research Triangle Park, NC, USA b EPL, Inc., Research Triangle Park, NC 27709-2194, USA c ExxonMobil Biomedical Services, Inc., Annandale, NJ, USA d Toxicology Consultants, Inc., Gibsonia, PA, USA Received 9 May 2003; received in revised form 2 October 2003; accepted 15 October 2003

Abstract Butyl benzyl phthalate (BBP) was administered in the diet at 0, 750, 3750, and 11,250 ppm ad libitum to 30 rats per sex per dose for two offspring generations, one litter/breeding pair/generation, through weaning of F2 litters. Adult F0 systemic toxicity and adult F1 systemic and reproductive toxicity were present at 11,250 ppm (750 mg/kg per day). At 11,250 ppm, there were reduced F1 and F2 male anogenital distance (AGD) and body weights/litter during lactation, delayed acquisition of puberty in F1 males and females, retention of nipples and areolae in F1 and F2 males, and male reproductive system malformations. At 3750 ppm (250 mg/kg per day), only reduced F1 and F2 offspring male AGD was present. There were no effects on parents or offspring at 750 ppm (50 mg/kg per day). The F1 parental systemic and reproductive toxicity no observable adverse effect level (NOAEL) was 3750 ppm. The offspring toxicity NOAEL was 3750 ppm. The offspring toxicity no observable effect level (NOEL) was 750 ppm, based on the presence of reduced AGD in F1 and F2 males at birth at 3750 ppm, but no effects on reproductive development, structures, or functions. © 2003 Elsevier Inc. All rights reserved. Keywords: Butyl benzyl phthalate; CAS No. 85-68-7; Dietary exposure; Adult systemic toxicity; Adult male reproductive toxicity; Offspring toxicity; Male offspring reproductive tract malformations; U.S. EPA OPPTS 837.3800 testing guideline

1. Introduction The phthalates, diesters of benzene-1,2-dicarboxylic (phthalic) acid, are used in vinyl tile, inks, solvents, and adhesives, as well as plasticizers for polyvinyl chloride (PVC) materials in a wide range of applications. The commercially important phthalates are generally linear or branched dialkyl esters of phthalic acid; aryl and saturated cyclic hydrocarbon side chains are also found. Because of their wide usage and applications, phthalates may enter the food chain and the environment. They are generally not persistent, and their potential for bioaccumulation is limited by biotransformation. The phthalates are metabolized in the gut in mammals to the monoester phthalate (e.g. butyl benzyl phthalate (BBP) is metabolized to monobenzyl phthalate (mBeP) and monobutyl phthalate (mBuP)). There is evidence that

∗ Corresponding author. Tel.: +1-919-541-5972; fax: +1-919-541-5956. E-mail address: [email protected] (R.W. Tyl).

0890-6238/$ – see front matter © 2003 Elsevier Inc. All rights reserved. doi:10.1016/j.reprotox.2003.10.006

the monoester phthalate metabolic product, not the parent diester phthalate, is the proximate toxicant. Sharpe et al. [1] reported that some phthalates (d-n-butyl phthalate (DBP) and BBP, but not diethylhexyl phthalate (DEHP)) had estrogenic activity, based on in vitro receptor binding assays or reporter cell systems [2,3]. Data indicating binding to the estrogen receptor (ER) and displacement of the endogenous ligand (17␤-estradiol (E2)) were reported for DEHP, BBP, DBP, and dihexylphthalate (DHP) at very high concentrations relative to the endogenous ligand E2. Studies with other in vitro systems (e.g. the Yeast Estrogen Screen (YES)) indicated that BBP, DBP, DEHP, and di-isobutyl phthalate (DiBP) had estrogenic activity (with di-isononyl phthalate (DiNP) exhibiting an equivocal response) again at very high doses, and with a response plateau at about 50% of that produced by E2, implying that they may be partial agonists. However, this mechanism is now considered unlikely [4,5] since none of the phthalates tested that were active in vitro displayed any estrogenic activity in vivo, such as uterotrophic growth response, accelerated acquisi-

242

R.W. Tyl et al. / Reproductive Toxicology 18 (2004) 241–264

tion of vaginal patency, induction of female sexual behavior, or vaginal cornification [6,7]. The current consensus is that the estrogenic activity of phthalates identified in the in vitro studies is “not relevant to humans or to the environment” [5]. There have been little or no human exposure data, but the presumption (based on data primarily from DEHP) has been that the levels of human exposure from manufacturing the phthalates, incorporating the phthalates into end-use products, use of the products, disposal of the products, and migration into the ecosystem are very low, so that the margin of exposure (MOE; ratio of the most sensitive animal NOAEL to human exposure levels from all sources) has been assumed to be relatively high. Blount et al. [8] recently analyzed urine samples from a CDC reference population of 289 adult humans for the monoester metabolites (to preclude external contamination) of seven commonly used phthalates, namely: monoethyl-, monobenzyl-, monobutyl-, cyclohexyl-, 2-ethylhexyl-, isononyl-, and octyl-phthalate. The data were presented as nanograms monoester per ml urine or per gram urinary creatinine. The monoesters with the highest urinary levels found were: monoethyl, monobutyl, and monobenzyl phthalate, presumably reflecting exposure to DEP (diethyl phthalate) and DBP and/or BBP. The 95th percentile at the urinary levels were 3750 ppb and 2610 ␮g/g creatinine for monoethyl phthalate, 294 ppb and 162 ␮g/g creatinine for monobutyl phthalate, and 137 ppb and 92 ␮g/g creatinine for monobenzyl phthalate. Women of reproductive age (20–40 years) had significantly higher levels of MBuP than other age/gender groups (P < 0.005); urinary levels of MBuP were subsequently reported by CDC to be similar or slightly lower in the most recent national report on Human Exposure to Environmental Chemicals [9,10]. The sources of exposure, especially to young women, are not known, although some personal care products do contain phthalates as “inert” excipients; phthalates are also used in food packaging and processing materials [11]. David [11] calculated the estimated intakes from the data by Blount et al. [8] and reported them to be within previously determined exposure estimates and below developed RfDs. Kohn et al. [12] also confirmed that previously calculated human daily intake estimates are in good agreement with the CDC median values. They did state “However, the maximal values of excreted monoesters indicate that some individual exposures are substantially higher than previously estimated for the general population.” Subsequent CDC biomonitoring results with larger human populations have failed to confirm such higher exposure values [9,10]. The National Toxicology Program (NTP) established the NTP Center for the Evaluation of Risks to Human Reproduction (CERHR) to provide scientifically based, uniform risk assessments of the evidence for reproductive and developmental toxicity of manmade or naturally occurring chemicals or chemical mixtures. The first panel was convened in 1999 to evaluate seven phthalates, including BBP. The CERHR phthalate reports were released for public com-

ment in October 2000 and have recently been published (BBP) [13]. In the CERHR BBP report, human exposure data were presented. The presumption was that exposure to the general population was almost exclusively in foodstuffs. They cited IPCS estimates of 2 ␮g/kg BW per day in adults, with exposures to children possibly up to three-fold higher. They also cited JMAFF (Japanese Ministry of Agriculture, Forestry and Fisheries) estimates of 0.11–0.29 ␮g/kg BW per day for adults and exposure to infants through formula at 0.1–02 ␮g/kg BW per day. The CERHR report concluded that BBP exposure to the general population, including children, was well below 10 ␮g/kg BW per day. Rodent reproductive and developmental toxicity studies have been reported on many phthalates, especially DEHP. There have been numerous studies of DEHP, focusing particularly on its potential to induce liver tumors in rats and mice. These tumors appear to be the consequence of peroxisome proliferating mechanisms (peroxisome proliferator-activated receptor alpha; PPAR␣), resulting in hepatomegaly and hepatocarcinogenesis in susceptible rodents. The possibility exists that PPAR␣ induction could also influence reproductive processes. However, knock-out mice (missing the PPAR␣), exposed to high oral doses of DEHP [14], exhibit the characteristic phthalate testicular toxicity (albeit diminished), so the reproductive toxicity is not predominantly mediated through peroxisome proliferation. Developmental toxicity studies have been completed on BBP [15,16] that employed maternal exposures on gestational day (GD) 6–15 during major organogenesis. These studies did not include dosing during the time of reproductive development for the “last trimester” of rodent gestation, and did not evaluate effects on puberty and adult reproductive structures and functions. Piersma et al. [17] did report on a developmental toxicity study of BBP in rats that included a dosing scheme for days 5–16 of gestation, as well as dosing for GD 5–20, and noted that the extended dosing regimen may be more sensitive. Recent data support the characterization of some phthalates as anti-androgens, acting through a mechanism that does not involve androgen receptor (AR) binding [18–24]. DBP (which shares a monoester metabolite, MBuP, with BBP) has been shown to disrupt normal male reproductive tract development, resulting in characteristic male reproductive tract malformations from perinatal gavage exposure to the dams [18–20,23]. The NOAEL for offspring toxicity was identified at 50 mg/kg per day for DBP by oral gavage [20] and at 330 mg/kg per day by dietary exposure [25]. Gray et al. [22,23] have reported that DEHP, DBP, BBP (with equivalent potency), and DiNP (approximately 10- to 20-fold less potent), but not DEP, DMP (dimethyl phthalate), or DOTP (di-octyl terphthalate), when administered at 750 mg/kg per day by gavage on GD 14 to postnatal day (PND) 3, result in retained nipples in preweanling male offspring and adult male offspring reproductive tract malformations. The mechanism of this effect appears to act through reductions in fetal testicular testosterone (T) synthesis [21].

R.W. Tyl et al. / Reproductive Toxicology 18 (2004) 241–264

Ema et al. [26] have reported increased preimplantation loss in Wistar rats from BBP at ≥750 mg/kg per day when gavage exposure was on GD 0–8, and increased postimplantation loss from BBP at 1000 mg/kg per day when gavage exposure was on GD 15–17 [27]. Nagao et al. [28] dosed CD® (Sprague–Dawley (SD)) rats with BBP by gavage at 0, 20, 100, or 500 mg/kg per day over two offspring generations. They reported reduced anogenital distance (AGD) in offspring males, decreased testes and epididymides weights in males, and decreased ovarian and increased uterine weights in females at 500 mg/kg per day (but not at 20 or 100 mg/kg per day). They did not report any offspring male reproductive system malformations at any dose (see Section 4 for more information). Therefore, the present study was performed to evaluate the potential of BBP, administered in the feed to CD® (SD) rats, to produce alterations in parental fertility, maternal pregnancy and lactation, and growth and development of the offspring, especially in the reproductive system (including accessory sex organs and external genitalia) for two adult and two offspring generations, 1 l/breeding pair/generation. This study was performed under the U.S. EPA OPPTS testing guidelines [29] with the following enhancements: • AGD and body weight were recorded for all live F1 and F2 offspring at birth on PND 0 (the OPPTS guidelines have AGD “triggered” in F2 offspring if developmental and/or reproductive effects are observed in F1 offspring). • F1 and F2 litters were standardized to 10 (with as even a sex ratio as possible) on PND 4 to remove the confounder of litter size on offspring survival and growth during lactation. All culled pups on PND 4 were subjected to an external and visceral examination, with special attention on the male reproductive organs (not in the OPPTS guidelines). • All F1 and F2 male preweanling pups were examined on PND 11–13 for the presence of retained nipples and/or areolae (not in the OPPTS guidelines). • At weaning on PND 21, in addition to the three pups/sex/litter necropsied for F1 and F2 offspring and to the F1 offspring selected to parent the F2 generation, all nonselected pups were necropsied, with special attention on the male reproductive organs (not in the OPPTS guidelines). • Epididymal sperm number, motility, and morphology; enumeration of testicular homogenization-resistant spermatid heads for calculation of daily sperm production (DSP); and efficiency of DSP were assessed in all F0 and F1 adult males at scheduled necropsy. • Additional histopathology was performed on F0 and F1 adult males in all groups if they exhibited gross lesions, if they did not sire live litters (also F0 and F1 females if they did not produce live litters), and if there was evidence of potential treatment-related histopathologic findings in any organs at the high dose.

243

• The target dietary doses (0, 750, 3750, and 11,250 ppm) were selected to provide a BBP intake of approximately 750 mg/kg per day at the high dose, at which Gray et al. [23] reported very high incidences of male reproductive system malformations in rats from gavage exposure to the dam on GD 14 through PND 3. That dose acted as the “positive control.” The low dietary concentration (750 ppm) was chosen to provide a BBP intake of approximately 50 mg/kg per day, a dose identified as a NOEL in offspring rats from DBP (which shares a common metabolite with BBP, MbuP), administered by gavage to dams during gestation and lactation or during late gestation by Mylchreest et al. [18,19,30]. The mid dietary concentration (3750 ppm), providing a BBP intake of approximately 250 mg/kg per day, was selected to provide dose–response information on any affected endpoints. 2. Materials and methods 2.1. Test material Commercial BBP (Santicizer® 160 Plasticizer; CAS No. 85-68-7) was received from Solutia (Bridgeport, NJ) with purity of approximately 98.5% BBP. Rats were administered the test material in the diet (PMI Certified Rodent Diet No. 5002) at concentrations of 0, 750, 3750, or 11,250 ppm. Dietary concentrations for the two-generation study were selected as described in the Introduction. Dosed diet preparations were formulated by mixing BBP into the appropriate amount of diet for each dietary concentration. The feed was then mixed with blank feed in a V-Shell Blender (Lowe Industries, Inc., Crestwood, IL). Control diets were prepared in the same manner. Prior to the start of the study, stability of BBP in the test diets was confirmed under frozen storage and cageside conditions. Dosed feed formulations were made approximately every 4–6 weeks and stored frozen. Feed jars were changed weekly. Verification of dosage concentrations was performed by GC with flame ionization detection (FID). Analyses showed that the BBP in the diet was mixed homogeneously, was stable frozen for at least 54 days and for at least 12 days under cageside conditions, and was administered at the desired feed concentrations throughout the study. There was no BBP detected in the control formulations, with an estimated limit of detection of 35 ppm. 2.2. Experimental design Only significant components to the study design are detailed below. All other facets were in compliance with the U.S. EPA Office of Prevention, Pesticides and Toxic Substances (OPPTS), Health Effects Test Guidelines, OPPTS 870.3800, Reproduction and Fertility Effects [29], EPA Toxic Substances Control Act (TSCA), Good Laboratory Practice (GLP) Standards, and the NRC Guide for the Care and Use of Laboratory Animals [31].

244

R.W. Tyl et al. / Reproductive Toxicology 18 (2004) 241–264

Fig. 1. Study design.

A graphic representation of the study design is presented in Fig. 1. The dietary concentrations were 0, 750, 3750, and 11,250 ppm. A total of 240 (120 males and 120 females) CD rats (Charles River Laboratories, Raleigh, NC) were assigned to the study at the initiation of the F0 10-week prebreed exposure period. Each group consisted of 30 males and 30 females to yield at least 20 pregnant females per group at or near term. Clinical signs for toxicity, body weights, and feed consumption were monitored according to guidelines. For the last 3 weeks of the prebreed exposure period, vaginal smears for estrous cyclicity and normality were taken for all F0 females. Abnormal cycles were defined as one or more estrous cycles in the 21-day period with prolonged estrus (≥3 days) and/or prolonged metestrus or diestrus (≥4 days) within a given cycle. The animals were mated (1:1) within groups following the 10-week prebreed exposure, for a period of 14 days, with no change in mating partners. On PND 0 (date of birth), all live pups were counted, weighed individually, the AGD measured, sexed, and examined grossly. On PND 4, the size of each F1 litter was adjusted to 10 pups by eliminating extra pups by random selection to yield, as nearly as possible, five males and five females per litter. Culled PND four pups were necropsied, with attention to the reproductive system. On PND 11–13, all F1 males were examined for retained nipples and/or areolae. On PND 21, each litter was weaned, and at least one F1 male and one F1 female pup/litter, if possible, were randomly selected for 30 per sex per group to produce the F2 generation. There were 26–29 l per group, so additional F1 males and females were selected from the maximum number of litters to obtain 30 per sex per group; the 0 and 750 ppm groups each had 26 l, and there were 28 l

at 3750 ppm and 27 l at 11,250 ppm. Following this selection, three weanlings/sex/litter, if possible, were randomly selected for necropsy. Any remaining nonselected F1 weanlings were also necropsied, with attention to the reproductive system. Selected animals of the F1 generation were administered BBP in the diet at their respective formulations for 10 weeks and then mated to produce the F2 generation, following the same study design as described for the F0 generation to produce the F1 generation, including AGD on PND 0, examination of culled pups on PND 4, examination of F2 males on PND 11–13 for retained nipples/areolae, and necropsy of all available F2 pups at weaning on PND 21. The study ended with the weaning of the F2 litters. Selected F1 and F2 weanling animals, all F0 and F1 parental animals, and PND 4 and 21 culled F1 and F2 pups were subjected to a complete gross necropsy. The stage of estrus at necropsy was determined for all F0 and F1 females. For weanling animals, the brain, spleen, thymus, ovaries (two), uterus with cervix and vagina, testes (two), epididymides (two), and seminal vesicles (two) were weighed. For parental animals, the brain, liver, kidneys, adrenal glands, spleen, ovaries, uterus, testes, epididymides, seminal vesicles with coagulating glands, and the prostate were weighed. Specific attention was focused on the examination of the parental reproductive organs, including determining the weights of the reproductive organs and accessory organs, including the prostate, for all males and weights of the paired ovaries and uterus and ovarian primordial follicle counts for high dose and control F0 and F1 females. The OPPTS testing guideline [29] allows flexibility on how ovarian primodial follicle counts are

R.W. Tyl et al. / Reproductive Toxicology 18 (2004) 241–264

determined. For this study, both ovaries per F0 and F1 female (ten per group for high-dose and control groups) were fixed in 10% neutral buffered formalin. Each ovary was trimmed, and the inner third of each ovary was placed in a tissue cassette and embedded in paraffin. Ten ovarian step-sections per ovary were taken at least 100 ␮m apart. Each step-section was approximately 3–4 ␮m thick and stained with hematoxylin–eosin (H&E). Primordial follicles were counted from each step-section according to criteria previously described [32,33]. Primordial follicles consisted of a single oocyte or an oocyte surrounded by one layer of flattened to cuboidal granulosa cells. An assessment of growing, antral, and Graafian follicles was additionally evaluated for each ovary but not counted. At the time of F0 and F1 parental male sacrifice, testicular homogenization-resistant spermatid head count, calculation of daily sperm production, and efficiency of daily sperm production were determined from one frozen testis per male for all males. In addition, the number, motility, and morphology of sperm from one cauda epididymis were evaluated in these same animals. Epididymal sperm motility and number were determined using an HTM TOX-IVOS Automated Sperm Analysis System (Version 10.8S, Hamilton-Thorne Research, Beverly, MA). Epididymal sperm morphology (200 sperm per male, if possible) was examined manually. Histopathologic evaluation of the ovaries with oviducts (two), vagina, uterus with cervix, testis, epididymis, seminal vesicles with coagulating glands (two), prostate, adrenals (two), thyroid, liver, kidneys (two), and any gross lesions was conducted on the F0 and F1 parental animals from the high dose and control groups. Male reproductive organs, livers, and kidneys were also examined in the mid and low doses for F0 and F1 parental animals. 2.3. Statistical analyses The unit of comparison was the male, the female, the pregnant female, or the litter, as appropriate. Treatment groups were compared to the concurrent control group using either parametric ANOVA under the standard assumptions or robust regression method [34–36] that do not assume homogeneity of variance or normality. The homogeneity of variance assumption was examined via Levene’s test [37], which is much more robust to the underlying distribution of the data than the traditional Bartlett’s test. If Levene’s test indicated lack of homogeneity of variance (P < 0.05), robust regression methods were used to test all treatment effects. The robust regression methods use variance estimators that make no assumptions regarding homogeneity of variance or normality of the data. They were used to test for overall treatment group differences, followed by individual tests for exposed versus control group comparisons (via Wald Chi-square tests), if the overall treatment effect was significant. The presence of linear trends was analyzed by GLM procedures for homogenous data or by robust regression methods for nonhomogenous data [38–44]. Standard

245

ANOVA methods, as well as Levene’s test, are available in the GLM procedure of SAS® Release 6.12 [44], and the robust regression methods are available in the REGRESS procedure of SUDAAN® Release 7.5.3 [45]. If Levene’s test did not reject the hypothesis of homogeneous variances, standard ANOVA techniques were applied for comparing the treatment group values to the control group values. The GLM procedure in SAS® 6.12 was used to evaluate the overall effect of treatment and, when a significant treatment effect was present, to compare each exposed group to control via Dunnett’s test [46,47]. For the litter-derived percentage data (e.g. periodic pup survival indices), the data were subjected to arcsin of the square root transformation prior to analysis. The transformed data were analyzed using ANOVA, weighted for litter size. A one-tailed test (i.e. Dunnett’s test) was used for all pairwise comparisons to the vehicle control group, except that a two-tailed test was used for parental and pup body weight and organ weight parameters, feed consumption, percent males/litter, and anogenital distance. Frequency data such as reproductive indices (e.g. mating and fertility indices) were not transformed. All indices were analyzed by Chi-square test for Independence for differences among treatment groups [48] and by the Cochran–Armitage test for Linear Trend on Proportions [49–51]. When Chi-square revealed significant (P < 0.05) differences among groups, then a Fisher’s Exact probability test, with appropriate adjustments for multiple comparisons, was used for pairwise comparisons between each treatment group and the control group. Acquisition of developmental landmarks (e.g. vaginal patency and preputial separation), as well as F1 and F2 newborn anogenital distance, were analyzed by Analysis of Covariance (ANCOVA; in addition to ANOVA analysis) using body weight at acquisition or measurement as the covariate. For correlated data (e.g. body and organ weights at necropsy of weanlings, with more than one pup/sex/litter), SUDAAN® software [45] was used for analysis of overall significance, presence of trend, and pairwise comparisons to the control group values. SUDAAN® software, for analysis of correlated data, uses the individual animal (not the litter) as the “n” but incorporates the effects of intralitter clustering or correlation (i.e. the tendency of litter mates to respond similarly) by the use of generalized estimating equations (GEEs). A test for statistical outliers [41] was performed on parental body weights and feed consumption (in g per day) and on parental and weanling necropsy organ weights. If examination of pertinent study data did not provide a plausible biologically sound reason for inclusion of the data flagged as “outlier,” the data were excluded from summarization and analysis and were designated as outliers. If feed consumption data for a given animal for a given observational interval (e.g. study day (SD) 0–7, 7–14, 14–21, 21–28, 28–35, etc., during the prebreed exposure period) were designated outliers or “unrealistic” (e.g. negative value), then summarized data encompassing this period (e.g. SD 0–70

246

R.W. Tyl et al. / Reproductive Toxicology 18 (2004) 241–264

for the prebreed exposure period) was not included in this value. For all statistical tests, the significance limit of 0.05 (oneor two-tailed) was used as the criterion for significance.

mately the same over time, but the adult body weights increase over four-fold over the 10 weeks (see Figs. 2A–D and 3A–D). 3.1. Parental systemic parameters

3. Results All BBP-containing diets were within 90–110% of target concentrations (with no BBP in the control diets; estimated limit of detection (ELD) = 35 ppm). BBP consumption at 750, 3750, and 11,250 ppm ranged from approximately 40 (end of prebreed) to 150 (last week of lactation), 180–760, and 590–2330 mg/kg per day, respectively, depending on the age (body weight) and sex of the animals and the phase of the study; e.g. consumption was highest for dams during the last week of lactation, confounded by the pups self-feeding, and for the selected postweanlings early in the prebreed period (when they exhibit a major growth spurt and are eating more feed relative to their body weights than mature adults do), and lowest at the end of the prebreed period, especially for males, since the feed consumption is approxi-

During the prebreed period, F0 parental male body weights were reduced at 11,250 ppm only on SD 7. Weight change was reduced for SD 0–7 at 11,250 ppm and increased for SD 7–14 at 3750 ppm and for SD 63–70 at 3750 and 11,250 ppm. Overall during the prebreed period (SD 0–70), there were no effects on weight gain or feed consumption in g per day. Feed consumption in g/kg per day and food efficiency were variable. F0 male BBP intake exhibited the expected increases across dose groups and the expected decreases within groups over time (data not shown). In contrast, F1 male body weights were significantly reduced at 11,250 ppm throughout the entire prebreed (SD 0–70) and mating periods (SD 70–84) (Fig. 2A). Feed consumption in g per day was similarly reduced at 11,250 ppm throughout the prebreed period. Feed consumption in g/kg per day was increased at 11,250 ppm, and feed efficiency was essentially

Fig. 2. F1 Male and female body weights and BBP intake (data are presented as mean ± S.E.M.; ∗ P < 0.05). (A) F1 male body weights during prebreed and mating periods. (B) F1 male BBP intake during prebreed period. (C) F1 female body weights during prebreed period. (D) F1 female BBP intake during prebreed period.

R.W. Tyl et al. / Reproductive Toxicology 18 (2004) 241–264

247

Fig. 3. F1 maternal body weights and BBP intake during gestation and lactation (data are presented as mean ± S.E.M.; ∗ P < 0.05). (A) F1 female body weights during gestation. (B) F1 female BBP intake during gestation. (C) F1 female body weights during lactation. (D) F1 female BBP intake during lactation. Note: Maternal BBP intake during the last week of lactation is confounded by the pups self feeding.

unaffected. F1 male BBP intake (Fig. 2B) mirrored that of the F0 males. F0 and F1 females exhibited reduced body weights at 11,250 ppm throughout the prebreed period (Fig. 2C for F1 females; F0 female data not shown), with no effects on feed consumption or food efficiency. BBP intake exhibited the same pattern as in the F0 and F1 males (Fig. 2A and B for F1 females; F0 female data not shown). There were no treatment- or dose-related adverse clinical observations in either sex in either generation at any dietary dose (data not shown). F0 (data not shown) and F1 (Fig. 2A) females exhibited significant reductions in body weight throughout gestation (GD 0, 7, 14, and 20) at 11,250 ppm. BBP intake was comparable between generations and exhibited the expected increases across BBP groups and the expected slight decreases within groups across time for F0 (data not shown) and F1 (Fig. 3B) pregnant females. F0 (data not shown) and F1 (Fig. 3C) females exhibited significantly reduced body weights at 11,250 ppm during lactation on PND 0, 4, 7, and 14 (but not on PND 21 at weaning). BBP intake during lactation for F0 (data not shown) and F1 (Fig. 3D) females ex-

hibited the expected increase across groups and the larger increases in BBP intake in all groups for the last week of lactation (PND 14–21), due to the offspring self-feeding during this period. At necropsy, F0 and F1 males and females exhibited increased absolute and relative liver weights and increased relative liver weights in F1 males at 11,250 ppm (also increased absolute and relative liver weights in F1 males at 3750 ppm), with no effects on liver weights in F1 females, and histopathologic lesions in the liver at 11,250 ppm in F0 (but not F1) males and in F0 and F1 females (increased liver weights were also seen in F0 females at 3750 ppm with no hepatic histopathology; see below). Absolute and relative kidney weights were increased in F0 males at 11,250 ppm (with absolute but not relative kidney weights also increased at 3750 ppm in F0 males). No kidney weight changes were observed in F1 males at 11,250 ppm (but increased absolute and relative kidney weights were present in F1 males at 3750 ppm), increased relative kidney weight (but not absolute weight) in F0 females at 11,250 ppm (and absolute and relative kidney weights increased in F0

248

Table 1 Summary of F0 and F1 adult systemic toxicitya Parameter

BBP (ppm) F0 0

Males No. on study No. of deaths No. at necropsy

Organ weights Liver Ab Rc

599.25 ± 9.46

23.100 ± 0.493 3.856 ± 0.057

750

3750

30 1 29

30 0 30

583.79 ± 7.13

22.774 ± 0.555 3.898 ± 0.076

621.42 ± 10.67

24.834 ± 0.492 4.005 ± 0.063

11,250 30 1 29 586.81 ± 8.78

0 30 0 30 601.30 ± 9.12

750

3750

30 1 29

30 2 28

606.55 ± 10.10

26.197 ± 0.628∗∗∗ 4.460 ± 0.074∗∗∗

23.428 ± 0.517 3.890 ± 0.051

23.567 ± 0.531 3.888 ± 0.065

626.94 ± 9.64

25.873 ± 0.648∗∗ 4.118 ± 0.065∗

11,250 30 2 28 547.34 ± 11.01∗∗∗

24.567 ± 0.568 4.494 ± 0.067∗∗∗

Kidneys A R

4.348 ± 0.067 0.728 ± 0.010

4.217 ± 0.068 0.0724 ± 0.012

4.655 ± 0.066∗∗ 0.752 ± 0.010

4.683 ± 0.082∗∗ 0.800 ± 0.012∗∗∗

4.334 ± 0.082 0.723 ± 0.013

4.466 ± 0.072 0.738 ± 0.011

4.848 ± 0.077∗∗∗ 0.775 ± 0.011∗∗

4.150 ± 0.084 0.760 ± 0.011

Adrenal glands A R

0.0615 ± 0.0017 0.0103 ± 0.0003

0.0607 ± 0.0007 0.0104 ± 0.004

0.0603 ± 0.0021 0.0097 ± 0.0003

0.0607 ± 0.0016 0.0104 ± 0.0002

0.0583 ± 0.0018 0.0097 ± 0.0003

0.0594 ± 0.0017 0.0098 ± 0.0002

0.0600 ± 0.0017 0.0096 ± 0.0003

0.0584 ± 0.0013 0.0108 ± 0.0003∗

Brain A R

2.177 ± 0.014 0.365 ± 0.005

2.184 ± 0.016 0.375 ± 0.004

2.213 ± 0.011 0.359 ± 0.006

2.186 ± 0.015 0.374 ± 0.006

2.205 ± 0.016 0.369 ± 0.006

2.202 ± 0.017 0.365 ± 0.006

2.226 ± 0.013 0.357 ± 0.006

2.143 ± 0.023∗ 0.394 ± 0.006∗∗

Pancreas A R

0.909 ± 0.021 0.152 ± 0.004

0.860 ± 0.029 0.148 ± 0.005

0.849 ± 0.031 0.137 ± 0.004

0.890 ± 0.032 0.153 ± 0.006

0.753 ± 0.025 0.126 ± 0.004

0.824 ± 0.023 0.137 ± 0.004

0.856 ± 0.034∗ 0.137 ± 0.006

0.862 ± 0.035∗ 0.157 ± 0.005∗∗∗

Pituitary A R

0.0169 ± 0.0003 0.0028 ± 0.0000

0.0165 ± 0.0003 0.0028 ± 0.0001

0.0168 ± 0.0003 0.0027 ± 0.0000∗

0.0172 ± 0.0004 0.0029 ± 0.0001

0.0168 ± 0.0003 0.0028 ± 0.0001

0.0163 ± 0.0002 0.0027 ± 0.0000

0.0166 ± 0.0003 0.0026 ± 0.0000

0.0165 ± 0.0004 0.0030 ± 0.0000∗

0 0

0 0

0 0

0 8d

0 0

0 0

0 0

0 0

30 1 29

30 1 29

30 0 30

30 0 30

30 0 30

30 0 30

30 0 30

Gross findings Histopathology: liver Females No. on study No. of deaths No. at necropsy Necropsy Body weight

342.21 ± 3.76

342.04 ± 5.17

342.52 ± 3.71

30 0 30 330.08 ± 4.18

338.70 ± 5.00

342.17 ± 5.43

349.21 ± 3.97

319.11 ± 3.98∗∗

R.W. Tyl et al. / Reproductive Toxicology 18 (2004) 241–264

Necropsy Body weight

30 0 30

F1

Adult body weights and BBP intake are presented in the figures. Data for organs with no effects on absolute or relative weights (e.g. spleen and thyroid in males and adrenal glands, spleen, brain, thyroid, pancreas, and pituitary in females) in any generation are not presented. Other tables present reproductive toxicity parameters. b A: absolute organ weight in grams (g). c R: organ weight relative to terminal body weight in percentage. d The number of animals with histopathologic findings in the liver. ∗ P < 0.05 vs. control group value. ∗∗ P < 0.01 vs. control group value. ∗∗∗ P < 0.001 vs. control group value.

a

0 5 0 0 Gross findings Histopathology: liver

0 1d

0 0

0 2

0 9

0 0

0 0

2.904 ± 0.067 0.909 ± 0.016 3.191 ± 0.052∗∗∗ 0.915 ± 0.014 3.015 ± 0.048 0.883 ± 0.012 2.943 ± 0.040 0.873 ± 0.014 3.068 ± 0.057 0.927 ± 0.013∗∗∗ 3.09 ± 0.050∗ 0.902 ± 0.012∗ 2.907 ± 0.046 0.850 ± 0.012 Kidneys A R

2.953 ± 0.054 0.867 ± 0.007

18.598 ± 0.411 5.452 ± 0.115 18.348 ± 0.471 5.434 ± 0.139 17.871 ± 0.500 5.208 ± 0.115 Organ weights Liver A R

18.525 ± 0.494 5.438 ± 0.149

19.473 ± 0.512 5.679 ± 0.136

20.511 ± 0.775∗∗ 6.217 ± 0.201∗∗∗

19.766 ± 0.553 5.671 ± 0.153

18.397 ± 0.928 5.736 ± 0.264

R.W. Tyl et al. / Reproductive Toxicology 18 (2004) 241–264

249

females at 3750 ppm), with absolute kidney weights increased at 3750 ppm, and no effects on relative weights at any dietary concentration in F1 females. The relative and absolute and/or relative kidney weights were not accompanied by any evidence of treatment-related gross or histopathologic lesions in either sex, in either generation, in any dose group (Table 1). Since the phthalates as a class, and BBP specifically [52,53], are known inducers of proliferation of peroxisomes (membrane-bound organelles) in the rodent liver, the observations of increased absolute and relative liver weights and the presence of hepatic histopathologic findings from BBP exposure at 11,250 ppm were expected (incidences were 0, 0, 0, and 8 in F0 males; 0, 0, 0, and 0 in F1 males; 1, 0, 2, and 9 in F0 females; and 0, 0, 0, and 5 in F1 females at 0, 750, 3750, and 11,250 ppm, respectively). The histopathologic findings in the liver at 11,250 ppm were described by the study pathologist as “Cytologic Alteration, Hepatic, Minimal” which comprised a spectrum of subtle to slight histologic changes that included diffuse cytomegaly and variable karyomegaly, reduced cytoplasmic glycogen, increased cytoplasmic eosinophilia, and increased cytoplasmic granularity, with the eosinophilic granularity representing increased numbers of peroxisomes. Unequivocal evidence of treatment-related, increased incidence of hepatocellular cytologic alteration was not confirmed in the mid and low dose F0 and F1 males and females (Table 1). 3.2. Parental reproductive toxicity BBP affects males exposed during gestation. Therefore, the findings for F0 parental males (beginning exposure to BBP as postpubertal animals) and for F1 parental males (beginning exposure to BBP as gametes, with exposure through gestation) are very different. There were no effects on reproductive structures or functions in F0 males at any dietary dose. Also as expected, there were no effects in F0 females on any parameters of estrous cycling, mating, or gestation. The F0 females did exhibit reduced absolute and relative weights of the paired ovaries and uterus at necropsy at 11,250 ppm in the absence of any differences in ovarian primordial follicle count at this dose (Table 2). In contrast, F1 parental males at 11,250 ppm (exposed since they were gametes) exhibited reduced mating and fertility indices, reduced absolute (but not relative) testes, epididymides, seminal vesicles/coagulating gland weights, and reduced absolute and relative prostate weights. They exhibited reduced epididymal sperm number, motility, progressive motility, and increased gross and histopathologic findings in the testis and epididymis. Gross male reproductive tract malformations in F1 adults at 11,250 ppm included hypospadias, missing reproductive organ or portion(s) of organs (predominantly the epididymides, testes, and prostate), and abnormal reproductive organ size and/or shape. Histopathologic examination confirmed missing organs or parts of organs,

250

R.W. Tyl et al. / Reproductive Toxicology 18 (2004) 241–264

aspermia in the epididymides, testicular seminiferous tubule degeneration and atrophy, and dilatation of the rete testis at 11,250 ppm in F1 male adults (Table 2). F1 parental females at 11,250 ppm (exposed since they were gametes) exhibited reduced mating and fertility indices (most likely due to the F1 males which also exhibited reductions in these indices), reduced uterine implantation sites, and total and live pups per litter on PND 0 (with no increase in dead pups per litter). They also exhibited an increased incidence of fluid-filled uteri (since the affected females were in estrus at demise, the consequence, if any, is unknown) and increased absolute and relative uterine weight (most likely due to the retention of fluid at estrus in this organ), with no histopathologic lesions in female reproductive tract organs. Consistent with the F0 females, the F1 females also exhibited reduced ovarian weights at 11,250 ppm in the absence of any effects on ovarian primordial follicle counts at this dietary dose (Table 2). F1 females carrying F2 litters at 11,250 ppm exhibited reduced number of litters (17 ppm versus 29, 28, and 26 at 0, 750, and 3750 ppm, respectively) and reduced number of total and live pups/litter at birth, with no effects on pre- or postnatal survival (Table 3). 3.3. Offspring systemic and reproductive toxicity Summaries of F1 and F2 offspring systemic toxicity are presented in Tables 3 and 4 and Fig. 4. Body weights per litter (sexes combined) of F1 (Fig. 4A) and F2 (Fig. 4B) offspring during lactation exhibited significant reductions at the high dose. For F1 offspring, the body weights were reduced on PND 0, 4, 7, 14, and 21, i.e. throughout lactation (Fig. 4A), and for F2 offspring, the body weights were reduced on PND 7, 14, and 21 (but not on PND 0 or 4) (Fig. 4B). At the necropsy of the F1 PND 4 culls (involving 48, 38, 51, and 45 F1 males and 55, 37, 45, and 30 F1 females at 0, 750, 3750, and 11,250 ppm, respectively), two males at 11,250 ppm exhibited undescended testes. At necropsy of the F1 weanlings (Table 4), both F1 males and F1 females at 11,250 ppm exhibited reduced terminal body weights, reduced absolute (but not relative) thymus weights, reduced absolute and relative spleen weights, and reduced absolute and increased relative brain weights (F1 males also exhibited increased absolute brain weights at 3750 ppm). F1 male weanlings also exhibited reduced absolute and relative testes weights at 11,250 ppm (and increased absolute and relative testes weight at 3750 ppm), and decreased absolute epididymal weight, with relative epididymal weight unaffected. F1 weanling females exhibited reduced absolute ovarian and uterine weights, with relative weights of both organs unaffected. F1 males also exhibited gross reproductive organ malformations at 11,250 ppm, including missing epididymis (whole or part), small epididymis(ides) missing, small testis, and undescended testis(es), with 25 F1 males affected (32.9%) out of 76 F1 males, including eight F1 weanling males with uni- or bilateral undescended testes.

There were no gross findings in any other organ examined in either sex at any dose (Table 4). At the necropsy of the PND 4 F2 culled pups (involving 61, 61, 47, and 14 males and 55, 47, 56, and 15 females at 0, 750, 3750, and 11,250 ppm, respectively), one male at 11,250 ppm exhibited hypospadias. At the F2 weanling necropsy (Table 4), both F2 males and females at 11,250 ppm exhibited reduced terminal body weights, reduced absolute (but not relative) thymus weights, reduced absolute and relative spleen weights, and increased relative (but not absolute) brain weights. The F2 males also exhibited reduced absolute and relative testes weights, with no effects on absolute or relative epididymal weights and increased incidence of gross findings in the male reproductive organs, all at 11,250 ppm only. At 11,250 ppm, 13 of 54 F2 male weanlings (24.17%) exhibited gross malformations (plus seven with nipples only), with a similar profile to that of the F1 male weanlings at this dose (see above). Of the 13 affected males at 11,250 ppm, five exhibited missing uni- or bilateral seminal vesicles, and 12 (including the five above) exhibited missing uni- or bilateral epididymis(mides), whole or part. One male at 750 ppm exhibited missing right testis and epididymis. There were no undescended testes found on PND 4 or 21 for F2 males. F2 female weanlings exhibited reduced absolute (but not relative) ovarian weights at 11,250 ppm and increased absolute (but not relative) uterine weight at 3750 ppm, with no effects on uterine weight at 11,250 ppm. There were no treatment-related gross findings in the female weanlings (Table 4). There were no postweaning evaluations since the study was completed at the weaning and termination of the F2 litters. As expected, the gestational, birth, and lactational survival indices of the F1 litters (generated from F0 parents, which began their exposure after they had acquired puberty) were unaffected. The number of F1 litters on PND 0 were 26, 27, 28, and 27 at 0, 750, 3750, and 11,250 ppm, respectively. F1 pup pre- and postnatal survival was unaffected. F1 pup body weights per litter were reduced at 11,250 ppm on PND 0, 4, 7, 14, and 21 (and increased at 3750 ppm on PND 7) (Fig. 4A). However, the F1 male pups (exposed beginning as gametes through their gestation) did exhibit reduced anogenital distance in a dose-related manner at 3750 and 11,250 ppm. As expected, F1 female pup anogenital distance was unaffected (Table 3). F1 high-dose preweanling male offspring also exhibited increases in the incidence (number of pups) and severity (number per pup) of retained nipples and/or areolae (Table 3). F2 pup body weights/litter at 11,250 ppm were reduced on PND 7, 14, and 21 (but not on PND 0 or 4) (Fig. 4B). F2 male AGD was significantly reduced at birth at 3750 and 11,250 ppm, with no effects on F2 female AGD. F2 male nipple and/or areolae retention was also increased at 11,250 ppm (Table 3). The F1 offspring exhibited significant increases in male reproductive tract malformations at 11,250 ppm on PND 21 (at weaning) and at adult necropsy. Two culled F1 males on

Table 2 Summary of parental F0 and F1 reproductive toxicity Parameter

BBP (ppm) F0

Adult male necropsy No. of males Terminal body weight (g) Organ weights Paired testes Ab Rc

0

750

3750

11,250

0

750

3750

11,250

30 100.0 90.0 96.3 90.0 3.1 ± 0.6a 22.2 ± 0.1 29 (96.7) 6 (20.7)

30 96.7 96.6 96.4 100.0 2.5 ± 0.5 22.2 ± 0.1 28 (93.3) 7 (25.0)

30 100.0 93.3 100.0 93.3 2.9 ± 0.4 22.0 ± 0.1 30 (100.0) 3 (10.0)

30 100.0 93.3 96.4 93.3 3.1 ± 0.6 22.0 ± 0.1 30 (100.0) 6 (20.0)

30 96.7 100.0 100.0 100.0 2.8 ± 0.2 22.0 ± 0.1 30 (100.0) 5 (16.7)

30 96.7 96.6 100.0 96.6 2.7 ± 0.4 22.1 ± 0.1 29 (96.7) 2 (6.9)

30 93.3 92.9 100.0 96.3 2.7 ± 0.3 22.0 ± 0.0 29 (96.7) 9 (31.0)

30 70.0* 81.0* 100.0 85.0 2.9 ± 0.6 22.0 ± 0.1 30 (100.0) 2 (6.7)

4.39 ± 0.20

4.58 ± 0.33

30 599.25 ± 9.46

30 583.79 ± 7.13

4.81 ± 0.24 30 621.42 ± 10.67

4.39 ± 0.11

4.38 ± 0.10

30 586.81 ± 8.78

30 601.30 ± 9.12

4.27 ± 0.11 29 606.55 ± 10.10

4.75 ± 0.15 28 626.94 ± 9.64

4.55 ± 0.14 28 547.34 ± 11.01∗∗∗

3.447 ± 0.058 0.578 ± 0.010

3.555 ± 0.051 0.611 ± 0.011

3.062 ± 0.072 0.587 ± 0.014

3.570 ± 0.064 0.611 ± 0.013

3.598 ± 0.050 0.600 ± 0.010

3.649 ± 0.047 0.606 ± 0.012

3.623 ± 0.114 0.583 ± 0.022

2.858 ± 0.179∗∗∗ 0.521 ± 0.030

Paired epididymides A R

1.424 ± 0.019 0.238 ± 0.004

1.424 ± 0.018 0.245 ± 0.003

1.482 ± 0.020 0.240 ± 0.004

1.424 ± 0.018 0.244 ± 0.005

1.351 ± 0.028 0.226 ± 0.006

1.379 ± 0.017 0.230 ± 0.004

1.385 ± 0.037 0.223 ± 0.007

1.208 ± 0.054∗ 0.217 ± 0.009

Prostate A R

0.846 ± 0.034 0.142 ± 0.006

0.834 ± 0.042 0.144 ± 0.008

0.824 ± 0.039 0.134 ± 0.007

0.821 ± 0.039 0.142 ± 0.005

0.756 ± 0.037 0.126 ± 0.006

0.665 ± 0.033 0.110 ± 0.006

0.704 ± 0.030 0.113 ± 0.005

0.563 ± 0.029∗∗∗ 0.103 ± 0.005∗∗

Seminal vesicles with coagulating glands A 2.362 ± 0.059 R 0.396 ± 0.010

2.313 ± 0.056 0.398 ± 0.011

2.509 ± 0.049 0.405 ± 0.008

2.303 ± 0.063 0.396 ± 0.013

2.145 ± 0.052 0.358 ± 0.009

2.158 ± 0.050 0.357 ± 0.008

2.129 ± 0.052 0.341 ± 0.009

1.752 ± 0.094∗∗∗ 0.316 ± 0.016

946.71 ± 20.94 77.6 ± 1.3 64.6 ± 1.6 104.23 39.25 22.61 2.12

± ± ± ±

7.22 3.06 2.57 0.21

913.28 ± 24.81 78.7 ± 1.0 67.0 ± 1.8 127.11 49.25 27.57 2.17

± ± ± ±

5.51 2.22∗ 1.20 0.27

948.18 ± 24.26 77.5 ± 1.2 64.0 ± 1.7 129.47 50.90 28.09 2.33

± ± ± ±

6.33 2.71∗ 1.37 0.25

930.47 ± 21.23 78.5 ± 1.4 64.7 ± 2.0 121.07 47.25 26.26 2.11

± ± ± ±

8.70 3.66 1.89 0.26

825.59 ± 38.16 68.6 ± 4.0 57.3 ± 3.8 47.22 18.18 10.24 5.99

± ± ± ±

6.12 2.32 1.33 3.18

860.78 ± 25.91 74.0 ± 1.7 61.2 ± 2.1 53.62 21.39 11.63 2.77

± ± ± ±

6.66 2.77 1.45 0.31

845.35 ± 40.82 71.7 ± 3.7 60.1 ± 3.7 47.01 19.13 10.20 5.61

± ± ± ±

5.82 2.28 1.26 3.08

649.51 ± 70.25∗ 52.1 ± 6.4∗ 42.1 ± 5.7∗ 30.30 11.96 6.57 3.27

± ± ± ±

6.54 2.99 1.42 0.63

251

Andrology Epididymal sperm concentration (106 /g) % Motile sperm % Progressively motile sperm TSHCd DSPe Efficiency of DSP % Abnormal spermf

R.W. Tyl et al. / Reproductive Toxicology 18 (2004) 241–264

No. of mating pairs Mating index (%) Fertility index (%) Gestational index (%) Pregnancy index (%) Precoital interval (days) Gestational length (days) No. (%) of females cycling No. (%) of females with abnormal cycles Estrous cycle length (days)

F1

252

Table 2 (Continued ) Parameter

BBP (ppm) F0 0

Histopathology Testes Epididymides Liver

Organ weights Paired ovaries A R Uterus A R Ovarian follicle counth Stage of estrus at necropsyi Gross: fluid-filled uterus Histopathology (liver)

750 1g 0 0

30 342.21 ± 3.76

1 0 0 30 342.04 ± 5.17

3750

11,250

0 0 0

1 0 8

30 342.52 ± 3.71

30 330.08 ± 4.18

0

750 3 2 0

30 338.70 ± 5.00

0 0 0 30 342.17 ± 5.43

3750 4 3 0 30 349.21 ± 3.97

11,250 23 15 0 30 319.11 ± 3.98∗∗

0.153 ± 0.007 0.0495 ± 0.0002

0.165 ± 0.004 0.0480 ± 0.0013

0.157 ± 0.004 0.0460 ± 0.0009

0.133 ± 0.005∗∗ 0.0400 ± 0.0070∗

0.146 ± 0.004 0.043 ± 0.001

0.150 ± 0.005 0.044 ± 0.001

0.151 ± 0.004 0.043 ± 0.021

0.149 ± 0.005 0.047 ± 0.003∗

0.575 ± 0.021 0.169 ± 0.007

0.568 ± 0.025 0.167 ± 0.006

0.594 ± 0.003 0.174 ± 3.006

0.146 ± 0.020∗∗∗ 0.140 ± 0.009∗∗

0.570 ± 0.021 0.169 ± 0.006

0.570 ± 0.023 0.167 ± 0.007

0.622 ± 0.029 0.179 ± 0.009

0.685 ± 0.045∗ 0.217 ± 0.016∗∗

281.1 ± 34.1 NSD 0 1g

– NSD 1g 0

– NSD 0 2

272.9 ± 35.1 NSD 0 9

368.4 ± 26.3 NSD 0 0

– NSD 0 0

– NSD 1 0

414.9 ± 56.5 NSD 3 5

Data are presented as mean ± S.E.M. A: absolute organ weight in grams (g). c R: organ weight relative to terminal body weight in %. d TSHC: testicular homogenization resistant spermatid head counts in 106 /g testis. e DSP: daily sperm production in 106 /testis per day; efficiency of DSP is in 106 /g testis per day. f There were no statistically significant differences in any BBP-exposed group compared to the control group for “% abnormal sperm.” The mean values in the two F1 groups above 3.50% were due to four (at 0 ppm) and three (at 3750 ppm) males with increased numbers of malformed sperm. Of the four F1 males affected at 0 ppm, all sired live litters. Of the three F1 males affected at 3750 ppm, one sired a live litter. g The numbers indicate the number of animals observed with gross or histopathologic findings as indicated. h Ovarian primordial follicle counts were performed only on 10 control and 10 high dose F0 and F1 females. i Assessment of the stage of estrus at necropsy involved a vaginal smear for each female at termination and determination of the estrous stage. The numbers (and percent) of females in each of the four stages (proestrus, estrus, metestrus, and diestrus) in each dose group were calculated and the values statistically analyzed for differences in the number (or percent) of females in a particular stage across groups. There were no statistically or biologically significant differences in the proportion of females in any of the four stages for any of the dose groups. NSD: not significantly different. ∗ P < 0.05 vs. control group value. ∗∗ P < 0.01 vs. control group value. ∗∗∗ P < 0.001 vs. control group value. a

b

R.W. Tyl et al. / Reproductive Toxicology 18 (2004) 241–264

Adult female necropsy No. of females Terminal body weight (g)

F1

Table 3 Summary of F1 and F2 offspring toxicity Parameter

BBP (ppm) F1 0

750

26 26 15.33 ± 0.58a 15.79 ± 4.52 5.5 ± 3.8 94.5 ± 3.8 13.3 ± 0.9 97.2 ± 1.3 100.0 ± 0.0 99.6 ± 0.4 100.0 ± 0.0 99.6 ± 0.4

27 26 13.61 ± 0.80 17.65 ± 3.96 2.1 ± 1.0 97.9 ± 1.0 12.2 ± 0.9 96.3 ± 1.4 99.6 ± 0.5 99.3 ± 0.5 99.6 ± 0.4 98.5 ± 0.9

3750 28 28 14.54 ± 0.57 8.77 ± 1.99 1.0 ± 0.7 99.0 ± 0.7 13.4 ± 0.6 97.5 ± 1.0 99.3 ± 0.5 100.0 ± 0.0 99.6 ± 0.4 98.9 ± 0.6

11,250 27 27 13.79 ± 0.66 14.18 ± 3.96 3.9 ± 2.0 96.1 ± 2.0 13.1 ± 0.7 92.6 ± 2.7 99.3 ± 0.5 98.8 ± 0.6 99.2 ± 0.5 97.4 ± 1.0

0

750

3750

11,250

29 29 15.86 ± 0.34 10.02 ± 1.21 4.0 ± 1.1 96.0 ± 1.1 14.2 ± 0.3 98.3 ± 0.7 99.3 ± 0.5 99.7 ± 0.3 100.0 ± 0.0 99.0 ± 0.6

28 28 15.00 ± 0.56 8.75 ± 1.90 3.4 ± 1.2 96.6 ± 1.2 14.0 ± 0.5 98.1 ± 0.9 99.6 ± 0.4 99.3 ± 0.5 99.2 ± 0.5 98.2 ± 0.9

26 26 15.23 ± 0.60 6.67 ± 1.10 1.5 ± 0.5 98.5 ± 0.5 14.2 ± 0.6 96.9 ± 0.9 100.0 ± 0.0 100.0 ± 0.0 100.0 ± 0.0 100.0 ± 0.0

17 17 12.35 ± 0.83∗∗∗ 7.06 ± 2.36 1.5 ± 1.2 98.5 ± 1.2 11.4 ± 0.7∗∗∗ 95.4 ± 2.0 95.9 ± 4.1 96.9 ± 2.2 100.0 ± 0.0 94.1 ± 4.8

Anogenital distance/litter (PND 0; mm) Male Female

2.06 ± 0.03 0.96 ± 0.02

2.01 ± 0.04 0.97 ± 0.02

1.89 ± 0.02∗∗∗ 0.92 ± 0.02

1.71 ± 0.03∗∗∗ 0.92 ± 0.02

2.05 ± 0.01 0.98 ± 0.01

2.05 ± 0.02 0.97 ± 0.01

1.99 ± 0.01∗ 0.96 ± 0.01

1.77 ± 0.03∗∗∗ 0.99 ± 0.02

Body weight/litter (PND 0; g) Male Female

6.76 ± 0.10 6.35 ± 0.10

6.93 ± 0.13 6.67 ± 0.16

6.71 ± 0.11 6.41 ± 0.11

6.15 ± 0.11∗∗ 5.91 ± 0.11∗

6.63 ± 0.11 6.21 ± 0.10

6.67 ± 0.12 6.34 ± 0.11

6.50 ± 0.10 6.20 ± 0.08

6.29 ± 0.12 5.93 ± 0.09

Preweanling male pups with retained nipple areolae (PND 11–13) % Male pups with ≥1 nipple 0.00 No. of nipples per male 0.00 ± 0.00 % Male pups with ±1 areola 2.63 No. of areolae per male 0.07 ± 0.04

0.00 0.00 ± 0.00 0.00 0.00 ± 0.00

0.00 0.00 ± 0.00 0.76 0.02 ± 0.00

19.23∗∗∗ 0.72 ± 0.20∗∗∗ 32.3∗∗∗ 1.29 ± 0.33∗∗

0.00 0.00 ± 0.00 2.13 0.05 ± 0.03

0.00 0.00 ± 0.00 5.07 0.12 ± 0.04

0.00 0.00 ± 0.00 5.43 0.19 ± 0.08

40.9 ± 0.4 208.17 ± 2.28 41.0 ± 0.3

41.3 ± 0.4 217.23 ± 4.15 41.1 ± 0.3

40.6 ± 0.2 212.83 ± 2.46 40.6 ± 0.2

45.2 ± 0.4∗∗∗ 207.63 ± 4.00 45.4 ± 0.4∗∗∗

31.4 ± 0.3 108.25 ± 2.50 31.5 ± 0.3

32.0 ± 0.4 114.56 ± 3.41 31.6 ± 0.3

31.2 ± 0.3 109.10 ± 2.08 31.2 ± 0.3

34.1 ± 0.5∗∗∗ 106.30 ± 2.72 34.4 ± 0.3∗∗∗

Acquisition of puberty Males Age (days) at PPSc Weight at acquisition (g) Age (days) at PPS adjusted for body weight Females Age (days) at VPd Weight at acquisition (g) Age (days) at VP adjusted for body weight

Data are presented as mean ± S.E.M. Not statistically significantly different from control group value (large variance term). c PPS: preputial separation in males. d VP: vaginal patency in females. ∗ P < 0.05 vs. control group value. ∗∗ P < 0.01 vs. control group value. ∗∗∗ P < 0.001 vs. control group value.

16.46b 0.51 ± 0.24∗ 72.15∗∗∗ 3.14 ± 0.50∗∗∗

R.W. Tyl et al. / Reproductive Toxicology 18 (2004) 241–264

No. of live litters: (PND 0) (PND 21) No. of implantations/litter % Postimplantation loss/litter Stillbirth index Live birth index No. of live pups/litter (PND 0) 4-Day survival index 7-Day survival index 14-Day survival index 21-Day survival index Lactational index

F2

NOT DONE (F2 pups terminated at weaning; PND 21)

a

b

253

254

Table 4 Summary of F1 and F2 weanling necropsy results Parameter

BBP (ppm) F1

Males (n) Body weight (g)

0

750

3750

11,250

0

750

3750

11,250

68 49.45 ± 1.24a

68 49.57 ± 1.14

78 50.00 ± 0.76

76 40.35 ± 1.08∗∗∗

86 51.78 ± 0.90

83 51.91 ± 1.30

78 52.65 ± 0.97

54 45.89 ± 1.20∗∗∗

0.215 ± 0.007 0.435 ± 0.003

0.229 ± 0.009 0.460 ± 0.013

0.218 ± 0.007 0.436 ± 0.011

0.179 ± 0.007∗∗∗ 0.443 ± 0.022

0.236 ± 0.008 0.454 ± 0.009

0.230 ± 0.008 0.443 ± 0.012

0.224 ± 0.008 0.425 ± 0.012

0.204 ± 0.008∗∗ 0.444 ± 0.013

0.205 ± 0.009 0.412 ± 0.011

0.206 ± 0.008 0.415 ± 0.012

0.202 ± 0.007 0.401 ± 0.009

0.146 ± 0.005∗∗∗ 0.363 ± 0.009∗∗∗

0.211 ± 0.007 0.406 ± 0.009

0.220 ± 0.008 0.423 ± 0.011

0.214 ± 0.006 0.406 ± 0.009

0.155 ± 0.008∗∗∗ 0.334 ± 0.011∗∗∗

Brain A R

1.450 ± 0.010 2.987 ± 0.061

1.477 ± 0.011 3.024 ± 0.063

1.490 ± 0.010∗∗ 3.006 ± 0.044

1.406 ± 0.014∗ 3.551 ± 0.074∗∗∗

1.485 ± 0.012 2.896 ± 0.040

1.465 ± 0.014 2.786 ± 0.060

1.495 ± 0.012 2.867 ± 0.048

1.462 ± 0.014 3.223 ± 0.075∗∗∗

Paired testes A R

0.222 ± 0.006 0.450 ± 0.006

0.229 ± 0.006 0.464 ± 0.008

0.240 ± 0.003∗ 0.480 ± 0.006∗∗∗

0.165 ± 0.006∗∗∗ 0.404 ± 0.008∗∗∗

0.243 ± 0.008 0.468 ± 0.010

0.240 ± 0.008 0.460 ± 0.008

0.240 ± 0.005 0.458 ± 0.008

0.195 ± 0.005∗∗∗ 0.427 ± 0.009∗∗

Paired A

0.057 ± 0.002

0.056 ± 0.002

0.057 ± 0.002

0.044 ± 0.002∗∗∗

0.057 ± 0.002

0.053 ± 0.002

0.053 ± 0.002

0.050 ± 0.002

0.115 ± 0.003

0.113 ± 0.003

0.115 ± 0.003

0.111 ± 0.004

0.110 ± 0.003

0.103 ± 0.003

0.102 ± 0.004

0.108 ± 0.004

1 (1.47) 0 (0.00)

2 (2.94) 0 (0.00)

0 (0.00) 0 (0.00)

1 (1.16) 0 (0.00)

1 (1.20) 1 (1.20)

7 (8.97) 0 (0.00)

Spleen A R

Epididymides R No. (%) with ≥1 No. (%) with ≥1 RTMe

UTMd

Females (n) Body weight (g) Organ weights Thymus Ab Rc

73 47.86 ± 1.10a

68 49.01 ± 1.55

81 47.72 ± 0.87

2 (2.63) 25 (32.89)∗∗∗ 67 37.33 ± 1.15∗∗∗

87 48.45 ± 0.87

83 49.67 ± 0.87

77 49.96 ± 0.90

1 (1.85) 13 (24.07)∗∗ 43 42.99 ± 1.08∗∗∗

0.232 ± 0.008 0.485 ± 0.013

0.237 ± 0.010 0.483 ± 0.014

0.226 ± 0.007 0.473 ± 0.011

0.180 ± 0.008∗∗∗ 0.477 ± 0.013

0.239 ± 0.008 0.493 ± 0.010

0.232 ± 0.006 0.468 ± 0.009

0.226 ± 0.007 0.452 ± 0.011

0.204 ± 0.008∗∗∗ 0.472 ± 0.012

Spleen A R

0.200 ± 0.009 0.414 ± 0.011

0.217 ± 0.012 0.438 ± 0.015

0.191 ± 0.006 0.399 ± 0.009

0.132 ± 0.005∗∗∗ 0.354 ± 0.009∗∗∗

0.197 ± 0.006 0.404 ± 0.008

0.212 ± 0.006 0.427 ± 0.008∗

0.204 ± 0.005 0.408 ± 0.008

0.146 ± 0.007∗∗∗ 0.336 ± 0.010∗∗∗

Brain A R

1.419 ± 0.012 3.007 ± 0.056

1.437 ± 0.013 2.997 ± 0.084

1.436 ± 0.010 3.044 ± 0.046

1.360 ± 0.019∗∗∗ 3.722 ± 0.102∗∗∗

1.430 ± 0.013 2.978 ± 0.044

1.430 ± 0.010 2.909 ± 0.046

1.449 ± 0.012 2.925 ± 0.045

1.409 ± 0.018 3.317 ± 0.077∗∗∗

R.W. Tyl et al. / Reproductive Toxicology 18 (2004) 241–264

Organ weights Thymus Ab Rc

F2

Data presented as mean ± S.E.M. A: absolute organ weight in grams (g). c Relative organ weight (as % of necropsy body weight). d UTM: urinary tract malformations, including hydroureter, hydronephrosis, malformed shape of urinary bladder, and/or thinned urinary bladder wall. e RTM: reproductive tract malformations, including missing epididymis (whole or part), epididymis reduced in size, missing testes, testes reduced in size, and undescended testis(es) for males. ∗ P < 0.05, statistically significantly different from control value. ∗∗ P < 0.01, statistically significantly different from control value. ∗∗∗ P < 0.001, statistically significantly different from control value. b

a

1 (1.30) 0 (0.00) 1 (1.20) 0 (0.00) 0 (0.00) 0 (0.00) 0 (0.00) 0 (0.00) 1 (1.47) 0 (0.00) No. (%) with ≥1 UTMd No. (%) with ≥1 RTMe

0 (0.00) 0 (0.00)

1 (1.15) 0 (0.00)

0 (0.00) 0 (0.00)

0.069 ± 0.005 0.161 ± 0.010 0.093 ± 0.004∗ 0.186 ± 0.008 0.084 ± 0.005 0.170 ± 0.009 0.080 ± 0.004 0.164 ± 0.007 0.112 ± 0.005 0.234 ± 0.008 0.106 ± 0.007 0.215 ± 0.012 0.108 ± 0.006 0.224 ± 0.010 Uterus A R

0.086 ± 0.006∗∗ 0.231 ± 0.014

0.026 ± 0.001∗∗ 0.061 ± 0.003 0.030 ± 0.001 0.062 ± 0.002 0.032 ± 0.001 0.066 ± 0.002 Paired ovaries A R

0.035 ± 0.001 0.073 ± 0.002 0.036 ± 0.002 0.074 ± 0.002 0.034 ± 0.001 0.071 ± 0.002

0.026 ± 0.001∗∗∗ 0.068 ± 0.002

0.032 ± 0.001 0.065 ± 0.002

R.W. Tyl et al. / Reproductive Toxicology 18 (2004) 241–264

255

PND 4 (at 11,250 ppm) also exhibited a reproductive tract malformation (undescended testes; observed rarely or never in control males) (Table 4 and Fig. 5A). The F2 offspring (raised until weaning) also exhibited significant increases in male reproductive tract malformations at 11,250 ppm on PND 21 (Table 4 and Fig. 5B), with one culled F2 male on PND 4 at 11,250 ppm exhibiting hypospadias. There were no treatment-related increases in male reproductive tract malformations at 750 or 3750 ppm in either generation at any age evaluated. There were no effects on F1 or F2 female reproductive tracts. The incidence of urinary tract malformations was also examined in F1 and F2 males and females, with no treatment-related changes observed in either sex in either offspring generation at any dietary dose (Table 4). 4. Discussion The present study evaluated exposure of CD® (SD) rats to BBP in the diet ad libitum at 750, 3750, or 11,250 ppm for two generations, one litter/generation, with approximate BBP intakes of 0, 50, 250, or 750 mg/kg per day, respectively. 4.1. F0 and F1 systemic toxicity The top dietary concentration (11,250 ppm) resulted in systemic toxicity to both parental sexes in both generations. The toxicity was expressed as reduced body weights and weight changes, increased absolute and relative liver weights, accompanied by histopathologic findings in the liver, consistent with hepatomegaly (from hepatocyte hypertrophy) due to induction of peroxisome proliferation from exposure to BBP and/or from induction of metabolizing enzymes [54]. BBP has been shown to be a weak peroxisome proliferator, even in comparison to other phthalates [52,53]. Increased absolute and/or relative kidney weights in both sexes in both generations (also at 11,250 ppm) were not accompanied by any evidence of treatment-related renal histopathologic lesions in either sex or generation in any dose group. 4.2. F0 and F1 reproductive toxicity There was no evidence of reproductive toxicity in F0 males at any dose. Rats (and mice) are robust reproducers, with excess sperm capacity and the ability to reproduce normally in the presence of significant testicular damage. Therefore, reproductive performance, per se, is not a sensitive endpoint. However, the use of reproductive organ weights, extensive histopathology, andrologic assessments (epididymal sperm number, motility, morphology; testicular homogenization-resistant spermatid head counts, the calculation of daily sperm production (DSP) and efficiency of DSP), and estrous assessments provide

256

R.W. Tyl et al. / Reproductive Toxicology 18 (2004) 241–264

Fig. 4. (A) F1 and (B) F2 pup body weights per litter (sexes combined) during lactation (data are presented as mean ± S.E.M.; ∗ P < 0.05).

greatly increased sensitivity to detect reproductive effects. F0 females exhibited reduced absolute and relative paired ovary and uterine weights at 11,250 ppm. There were no effects of treatment on mating, fertility, or gestational outcome indices, in estrous cyclicity, or in ovarian primordial follicle counts. Pre- and perinatal loss, F1 litter size, and sex ratio of offspring were equivalent across groups. F0 male andrological parameters, reproductive organ weights, and histopathology were all unaffected at any dose. F0 female reproductive organ gross examination and histopathology were uneventful. None of these findings (or lack thereof) were unexpected since the F0 parents began exposure as postpubertal animals, after the reproductive structures and functions were in place and fully developed. The F1 offspring were exposed since they were gametes in their F0 parents, through gestation and lactation, and puberty to adulthood and reproduction to produce F2 litters. The F2 offspring were also similarly exposed from gametogenesis, through gestation and lactation, until scheduled demise at weaning on PND 21. Therefore, there were

profound effects on F1 and F2 male reproductive development (typical of in utero exposure to an anti-androgen). They included effects at 11,250 ppm on AGD (shortened), retention of nipples and/or areolae (increased), and acquisition of puberty (delayed). Adult F1 males at 11,250 ppm also exhibited reduced reproductive organ weights (paired testes, paired epididymides, prostate, and paired seminal vesicles with coagulating glands) and incidence of gross and histopathologic malformations of the male reproductive tract. F1 male andrology at 11,250 ppm was also affected, with reductions in epididymal sperm count, motility, and progressive motility. F1 females at 11,250 ppm exhibited no effects on AGD but did exhibit reduced paired absolute and relative ovarian weights and increased absolute and relative uterine weights (which is different from the F0 reduced absolute and relative uterine weights), with no reproductive organ histopathologic lesions. There were reduced numbers of F2 (but not F1) uterine implantation sites/dam and, therefore, reduced number of total and live F2 pups/litter at birth at 11,250 ppm. The only effect observed at 3750 ppm was significantly reduced AGD

R.W. Tyl et al. / Reproductive Toxicology 18 (2004) 241–264

257

offspring generations (shortened in a dose-related pattern), at a dose in the presence of reduced body weight at birth (11,250 ppm for F1 males), and at doses in the absence of reduced body weights at birth (11,250 ppm for F2 males and 3750 ppm for F1 and F2 males), support the interpretation of treatment-related effects on this parameter at these doses. However, the biological significance of this finding at 3750 ppm is unclear due to the absence of any effects on male reproductive development, structures, or functions at this dose. Historical control values in the performing laboratory for F2 male AGD are 1.96–2.25 mm. The F1 and F2 male mean AGDs at 3750 ppm in this study were 1.89 and 1.99 mm, respectively (F2 mean within historical range, F1 mean below historical range). The male offspring AGD data for this study are presented in Table 3. Gray et al. [6,23,59] and Parks et al. [60] have stated that AGD at birth and retention of nipples/areolae in preweaning males “are the most sensitive indicators of anti-androgenic activity.” However, there are three additional considerations:

Fig. 5. Incidence of (A) F1 and (B) F2 offspring male reproductive tract malformations (data are presented as mean ± S.E.M.; ∗ P < 0.05).

in F1 and F2 males at birth. Neither the F1 nor F2 litters exhibited any treatment-related decreases in postnatal survival at any dose. Although measurement of AGD is a “triggered” endpoint to be performed in F2 newborns if there are effects on reproductive development in F1 offspring, AGD was measured in both F1 and F2 offspring in the present study. The measurement in F1 offspring allowed the full range of parameters to be evaluated from AGD at birth, through sexual maturation (acquisition of puberty), adult mating, gestation and lactation of F2 offspring, organ weights, gross and histopathology of reproductive organs, andrology, and estrous cyclicity on the same F1 generation. Although pup body weights confound AGD (larger pups have longer distances, and smaller pups have shorter distances [55]), AGD is under dihydrotestosterone (DHT) control [56,57] and may be the most sensitive parameter in animals exposed in utero to an anti-androgen [58]. The findings in this study, that male AGD was affected in both

1. There were no consequences to the reduced F1 male AGD at 3750 ppm in terms of reproductive function (mating, fertility and pregnancy indices, precoital interval, gestational length, litter size, sex ratio, postnatal pup survival, retained nipples/areolae, etc.), reproductive organ weights, or histopathology. F2 pups were terminated at weaning on PND 21 (with no effects on retained nipples/areolae at 3750 ppm), but F1 pups were evaluated through reproduction and necropsied at adulthood. 2. McIntyre et al. [58,61] examined the predictive value of male AGD and male retained nipples/areolae for the presence of male reproductive system malformations as adults. They used flutamide [58] and linuron [61], both anti-androgens, and concluded that, although these early effects “may be suggestive of altered T-mediated reproductive development seen in adult rats, these endpoints are not predictive” [61]. 3. Bowman et al. [62] administered finasteride (a specific inhibitor of type II 5␣-reductase that converts T to DHT) by gavage to pregnant SD rats on GD 12 to 21 at 0, 0.01, 0.1, 1, 10, or 100 mg/kg per day. They reported dose-related effects in adult F1 males on AGD (reduced), retained areolae/nipples on PND 13, nipples on PND 90, and on incidences of hypospadias, ectopic testes, bulbourethral gland agenesis, and prostate agenesis. They performed a logistic regression of male reproductive tract malformations on PND 90 versus AGD at birth (designated PND 1) and presented predictions of what AGD (in mm) would result in specific malformations in 50% of the males. Although they did not provide a numerical value for control male AGD at birth, their Fig. 1 indicates a value of approximately 3.20 mm. No effects were predicted until the AGD was reduced to 2.82 mm (88.1% of the control value). At this reduction, they predicted areolae/nipples on PND 13. At 2.61 mm (81.6% of controls), they predicted nipples on PND 90, at 2.39 mm (74.7% of control)

258

R.W. Tyl et al. / Reproductive Toxicology 18 (2004) 241–264

they predicted hypospadias, at 2.16 mm (67.5% of controls) they predicted ectopic testes, at 2.13 mm (66.6% of controls) they predicted bulbourethral gland agenesis, and at 1.87 mm (58.4% of the control value) they predicted prostate agenesis. The last value is comparable to their control females’ AGD at birth of 1.83 ± 0.04 (S.E.) mm. They did indicate that some animals with reduced AGD and/or retained nipples “may not have had other reproductive tract malformations” (p. 393). They concluded “Therefore the relationship between AGD, nipple retention, and male reproductive tract malformations in adult rats following in utero exposure is strictly dependent on the potency of the anti-androgen and its mechanisms of action” [62] (p. 405). The three considerations are directly relevant to the conclusion of the present study that 3750 ppm, at which there were reduced F1 and F2 male AGDs at birth, is an effect level, but not an adverse effect level, since the male AGD at 3750 ppm was reduced only to 91.8% of control (F1) and to 97.1% of control (F2), and in agreement with Bowman et al.’s [62] predictions, there were no statistically or biologically significant increases in the incidence of nipples or areolae in preweanling males or in any treatment-related male reproductive system malformations on PND 4 or 21 (both generations) or in adulthood (F1 males) at this dietary dose in either generation. Even at 11,250 ppm in this study, the AGD reductions were to 83.0% of control for F1 males and 86.3% of control for F2 males at birth (the reduction at which only retained nipples were predicted by Bowman et al. [62] on PND 13 and as adults). AGD is therefore considered a “biomarker of exposure” at 3750 ppm and a reasonable predictor of adverse effects on the male reproductive system at a higher dose (i.e. 11,250 ppm), but not adverse, per se. Acquisition of puberty was significantly delayed in F1 offspring of both sexes at 11,250 ppm when statistically evaluated by ANOVA (analysis of variance) or by ANCOVA (analysis of covariance), with body weight at acquisition as the covariate. Acquisition of developmental landmarks is dependent on both age and weight, i.e. heavier animals acquire the landmark earlier, while lighter animals acquire the landmark later, but lighter animals do acquire the landmark (unless there is another cause for the delay) and in many cases acquire the landmark at a lighter weight than the heavier animals. In the present study, the body weights of the male and female F1 offspring at acquisition were not reduced at 11,250 ppm. However, the F1 animals at 11,250 ppm were older at acquisition than the animals at 0, 750, or 3750 ppm and were, therefore, comparable in body weight to the younger animals at acquisition in the lower dose groups and control group. In fact, body weights were reduced throughout the F1 postwean prebreed period in both sexes at this dose (Fig. 2B and C). Therefore, it is possible that reduced body weights resulted in or confounded the observed delay in acquisition of vaginal patency and preputial separation. The only endocrine-mediated mechanism cur-

rently known to result in delays in puberty in both sexes would be interference with steroidogenesis, thereby reducing T (and DHT) levels in males and estrogen levels in females, and some phthalates have been shown to interfere with fetal steroidogenesis. The U.S. EPA [63] has also recognized that “body weight at puberty may provide a means to separate specific delays in puberty from those that are related to general delays in development.” The delays in vaginal patency in females and in preputial separation in males at 11,250 ppm in this study were relatively minor in F1 females (2.7 days) and slightly longer in F1 males (4.3 days). The female delays are typical of those observed in the performing laboratory when systemic toxicity expressed as reduced body weights appears to be the sole cause [64,65]. The male delays may be due to both systemic toxicity and to the antiandrogenic effects of BBP. In addition, Kennedy and Mitra [66] showed that body weight and food intake are factors in the initiation of puberty in the rat. Carney et al. [67] fed SD pregnant dams at 70 or 50% of control feed intake levels from GD 7 through weaning of offspring. Feed restriction resulted in reduced body weights in dams and offspring and delayed vaginal patency and preputial separation (by 1 day at 70% and by 6 days at 50% restriction for both parameters). The most conservative interpretation for the delays in acquisition of puberty in the F1 males and females is that it may be due to effects on steroidogenesis (especially for the males), confounded by systemic toxicity in both sexes, at 11,250 ppm. The present study also examined the incidence of urinary tract and reproductive tract malformations in F1 and F2 male and female offspring on PND 4 and 21 and in F1 offspring as adults. The F1 males at 11,250 ppm exhibited a significant increase in reproductive tract malformations on PND 21 and as adults (with a slight but not statistically significant increase on PND 4, based on two F1 males with undescended testes) (Fig. 5A). Pnd 21 F1 males at 11,250 ppm also exhibited undescended testes. F1 males on PND 21 at 11,250 ppm also exhibited undescended testes. The F1 males did not exhibit any increases in urinary tract malformations. Gross urinary tract malformations included hydroureter, hydronephrosis, malformed urinary bladder and/or thinned wall of urinary bladder. F1 preweanlings exhibited increased incidences of the number of males with, and the number per male of, retained nipples and areolae at 11,250 ppm. As expected, F1 females did not exhibit any reproductive tract malformations and no treatment-related increases in urinary tract malformations, at any dose at any time. F2 litters, developing in F1 dams (sired by F1 males), exhibited reductions in the number of litters and the number of total and live pups per litter on PND 0 at 11,250 ppm. Survival during lactation was unaffected, as was sex ratio. Pup body weights were unaffected on PND 0 and 4 and reduced at 11,250 ppm for PND 7, 14, and 21. The F2 males also exhibited significant increases at 11,250 ppm in the percentage

R.W. Tyl et al. / Reproductive Toxicology 18 (2004) 241–264

of male pups on PND 11–13 with one or more nipples and with one or more areolae, and in the number of nipples and areolae per male. F2 males also exhibited a significant increase in male reproductive malformations on PND 21 (Fig. 5B) at 11,250 ppm, with no treatment-related increases in urinary tract malformations. Neither PND 4 culled nor PND 21 F2 males at 11,250 ppm exhibited undescended testes. As with the F1 litters, F2 females did not exhibit any reproductive tract malformations and no increases in the incidence of urinary tract malformations. Since the last step in the biosynthesis of 17␤-estradiol (E2; the mammalian endogenous estrogen) is conversion of T or androstenedione to E2 in the ovary, the effects on F0 and F1 adult female ovarian weights and F2 reduced litter sizes at birth (with reduced uterine implantation sites) at 11,250 ppm may have been due to diminished precursor androgen to produce E2 in the ovaries. The same argument (reduced T) can be made to explain reduced F1 and F2 offspring male AGD at 11,250 ppm (and at 3750 ppm; under DHT control), the increased incidence of undescended testes in PND 4 and 21 F1 male offspring, the increased retention of nipples and areolae in F1 and F2 preweanling males (regression of nipple anlagen is under DHT control), delays in acquisition of puberty in both male and female F1 postweanlings, and the increases in F1 and F2 male reproductive tract malformations (due to reduced systemic T) at 11,250 ppm. Shultz et al. [68,69] administered DBP at 500 mg/kg per day or flutamide at 50 mg/kg per day in corn oil to pregnant Sprague–Dawley rats by gavage on GD 12–21. Gene expression was examined on GD 16, 19, and 21 in control and exposed fetal testes using Clontech’s rat gene expression microassay. The expression levels of a number of genes were altered by at least two-fold by DBP, including a variety of genes involved in signal transduction, cell growth, apoptosis, and steroid biosynthesis. Most notably, P450-side chain cleavage (SCC) mRNA, for the enzyme which catalyzes the rate-limiting step in T biosynthesis, was decreased from DBP by nearly 70% on GD 16 (the time of peak T synthesis [70]) and was elevated approximately two-fold on GD 21. Flutamide exposure did not affect the expression levels of P450SCC. The DBP-induced elevation of P450 SCC mRNA correlated with a nearly 10-fold decline in T production on GD 19 and 21. In addition, a dramatic six-fold induction in T-repressed prostate message-2 (TRPM-2) mRNA was observed on GD 21, corresponding to the timing of reduced testicular T levels by DBP. Immunohistochemical analysis confirmed that TRPM-2 protein levels were similarly elevated in both Sertoli and interstitial Leydig cells on GD 21, suggesting its potential role in promoting cell survival in these cells. The authors concluded that these results “suggest that DBP toxicity may be mediated through an inhibition of T biosynthesis, resulting in the altered expression of T-repressed genes” [68]. Li et al. [71] have reported that a single gavage dose (of 20 ␮l/g) of DEHP (or MEHP but not 2-EH), to 3-day-old rat pups at 200 and 500 mg/kg per day (but not at 20 or 100 mg/kg per day), resulted in abnormal,

259

multinucleated, enlarged gonocytes which persisted through 48 h postdose (the latest timepoint evaluated), reduced Sertoli cell proliferation, and decreased cyclin D2 (cell cycle regulator) expression. Leydig cell status was not reported. For the evaluation of DEHP versus MEHP and 2-EH, the doses were 500 mg/kg per day DEHP, 393 mg/kg per day MEHP, and 167 mg/kg per day 2-EH (all at 1.28 mmol/kg). Imajima et al. [72] and Shono et al. [73] reported postnatal cryptorchidism, preceded by delays in the transabdominal descent of testes in rat fetuses, after gestational exposure to a very high oral dose (1000 mg/kg per day) of monobutyl phthalate (one of the two major intestinal metabolites of BBP) on GD 15–18 [63] or on GD 7–10, 11–14, or 15–18 [73]. Ema et al. [74] administered monobenzyl phthalate to pregnant Wistar rats by gavage at 0, 167, 250, or 375 mg/kg per day on GD 15–17 (16 dams per group) and evaluated the fetuses on GD 21. Maternal body weight gain and feed consumption were reduced at >167 mg/kg per day, and fetal body weight was reduced at 375 mg/kg per day. There were dose-related increases in undescended testes and dose-related decreases in AGD in male fetuses at 250 and 375 mg/kg per day (AGD of female fetuses was unaffected at any dose), so both of the monoester metabolites can cause these effects in male rat fetuses. Testes normally descend to the inguinal ring by term and into the scrotal sacs during late lactation (typically PND 16–20 in our laboratory). In this study, cryptorchidism or undescended testes was observed at a low incidence (in two F1 males at 11,250 ppm) at necropsy of the PND 4 male culls, with the undescended testes located in the mid abdomen instead of the control location in the lower abdomen, close to the inguinal ring on PND 4. Eight F1 males at 11,250 ppm on PND 21 exhibited one or both testes within the abdominal cavity at the inguinal ring, but not descended into the scrotal sacs. Control males have achieved testes descent into the scrotal sacs by PND 21. We did not see cryptorchid (undescended) testes in any F2 male evaluated in this study at any time or dose. The undescended testes on PND 4 and 21 in the F1 males in this study may have been due to delays in (not permanent failure of) testes descent, since adult F1 and F2 males did not exhibit undescended testes at any dose. The differences in early incidence and the absence of undescended testes in adult males in this study, versus those results of Imajima et al. [72], Shono et al. [73], and Ema et al. [74] are obviously the dose (1000 mg/kg per day by Imajima et al. and Shono et al., 167–375 mg/kg per day by Ema et al., and by 50–750 mg/kg per day in the present study), the moiety administered (MBuP or MBeP by them, BBP in the present study), and the route (gavage by them, dosed feed in the present study). The findings were qualitatively very similar; the fetal effects observed by Ema et al. at 250 and 375 mg/kg per day on testes descent may have been, in fact, transient. Very recent data have indicated that male perinatal reproductive development is regulated not just by T (and DHT) made by the Leydig cells and Müllerian-inhibiting substance (MIS; made by the Sertoli cells) which cause the Müllerian

260

R.W. Tyl et al. / Reproductive Toxicology 18 (2004) 241–264

ducts, which form female reproductive structures to regress in males, but also by a peptide hormone, now designated as insulin-like factor 3 (Insl3) produced by the Leydig cells [75]. Insl3 was discovered in the 1980s, but its function in regulating the development of the gubernaculum (which attaches to the caudal portion of the testis and epididymis and is responsible, in whole or part, for testis descent into the lower abdomen to the inguinal ring in utero and into the scrotal sacs during late lactation in rodents) was not identified until Insl3-knockout mice were constructed (with normal T and MIS), which exhibited bilateral cryptorchidism [76,77]. Insl3 has also been shown to play a role in DES-induced cryptorchidism in mice [78], which implies estrogen involvement. The role of Insl3 in human cryptorchidism is not clear, with evidence, both for its role [79] and against any role [80]. Emmen et al. [81] have shown that both androgen and Insl3 are required for rodent gubernuclear outgrowth in vitro. Gray et al. [23,82,83] have reported that gavage administration of BBP (or DEHP or DBP) at 750 mg/kg per day on GD 14 to PND 3 to rats results in male reproductive system malformations, including cryptorchidism associated with abnormalities of the gubernaculum. In addition, examination of fetal testes from rat dams exposed to DEHP at 750 mg/kg per day on GD 14–18 indicated that not only was T production reduced, but also that Insl3 mRNA was inhibited by approximately 80% [83]. When both DBP and BBP (each by gavage at 500 mg/kg per day, GD 14–18; Gray et al., in preparation) were administered simultaneously to rat dams, no offspring males in the combination group exhibited normal gubernacula, and a few males exhibited retained cranial suspensory ligaments, which normally maintain the ovaries in the upper abdomen in females. In some cases, only the gubernaculum was affected, with no other male reproductive system malformations or vice versa, indicating an effect of the phthalates at high oral doses on both T and Insl3 synthesis. In support of this interpretation, DBP administered daily by gavage at a high oral dose (500 mg/kg per day) to dams during gestation and lactation also produced a small but biologically significant incidence of undescended testes, as well as other male reproductive system malformations in male offspring on PND 21 (George et al., unpublished observations, 2003). It is likely that both mBeP and mBuP phthalates at high doses are male developmental toxicants in rats. Therefore, the cryptorchidism from administration of various phthalates (BBP, mBuP, mBeP, DBP, DEHP, etc.) observed by Imagima et al. [72], Shono et al. [73], Gray and Foster [82], Wilson et al. [83], George et al. (unpublished observations, 2003), and the present study resulted from very high oral doses during the sensitive period. The minimal presence of cryptorchidism only in young F1 males in the present study is likely due to the dietary route of administration, resulting in slow exposure over the feeding period, even at the same doses in mg/kg per day and during the same sensitive period of reproductive development, as employed by the other researchers listed above. The absence of cryptorchidism in F1 and F2 adult males may indicate that at

11,250 ppm, the slow dietary exposure caused only a delay in testes descent, not a permanent cryptorchidism. Nagao et al. [28] performed a two-generation study in Crj:CD(SD) IGS rats of BBP by oral gavage at 0, 20, 100, or 500 mg/kg per day. They reported reduced AGD in offspring males, decreased testes and epididymides weights, and decreased ovarian and increased uterine weights in females, all at 500 mg/kg per day, but not at 100 mg/kg per day. These data fit nicely with the present study data since we observed comparable effects at 750 mg/kg per day (11,250 ppm) but not at 250 mg/kg per day (3750 ppm). It is surprising that they did not report male reproductive system malformations at 500 mg/kg per day. They did report “. . . atrophy of the testis was observed in six males, atrophy of the epididymis in four males, and enlargement of the testis and atrophy of the prostate in one male each of the 500 mg/kg per day group.” These observations may, in fact, be malformations, i.e. not “atrophy” but agenesis, at 500 mg/kg per day. They did not report testicular cryptorchidism. Nagao et al. [28] also measured relevant hormone concentrations in the blood, with variable results. Ema and Miyawaki [27] recently reported a study in which pregnant Wistar rats (16 per group) were dosed with BBP by gavage at 0, 250, 500, or 1000 mg/kg per day on GD 15–17, and the dams and fetuses were terminated and examined on GD 21. Reduced maternal body weight gain and feed consumption were present at 500 and 1000 mg/kg per day. There were decreased numbers of live fetuses/litter (i.e. postimplantation loss) and reduced fetal body weights (males and females) per litter at 1000 mg/kg per day. Increased incidences of undescended testes and decreased AGD in the males were present at both 500 and 1000 mg/kg per day. Female fetal AGD was unaffected at any dose. In a developmental toxicity assessment of BBP administered by gavage in rats on GD 7–15, increased incidences of malformed fetuses and increased postimplantation loss were detected at 750 mg/kg per day [84]. The authors concluded that “the doses that produced impairment of the male reproductive system were lower than those that produced fetal malformations in other organ systems and postimplantation embryonic loss.” They therefore suggested that “the male reproductive system may be more susceptible than other organ systems to BBP toxicity after maternal exposure” [27]. Ema et al. [26] had previously reported increased preimplantation loss in Wistar rats exposed to BBP by gavage at ≥750 mg/kg per day on GD 0–8. The present study did not measure relevant hormone levels or employ a separate positive control chemical (neither required in the OPPTS testing guideline). However, hormone levels are supposed to change in response to a changing external (or internal) environment, with the complex positive and negative feedback loops among the hypothalamus, pituitary, and gonads (and adrenals, etc.). These delicate adjustments are termed “homeostasis” as the animal (or human) adjusts to external and internal changes. Changes in hormone levels, per se, are not adverse. It is the consequences,

R.W. Tyl et al. / Reproductive Toxicology 18 (2004) 241–264

if any, from the changes in hormone levels, which may or may not be adverse. For example, AGD is under the control of DHT, made from T by the enzyme 5␣-reductase in the Leydig cells of the testis. DHT is also responsible for the development of male-specific external genitalia and the prostate from the genital tubercle and urogenital sinus and regression of the nipples/areolae in fetal males [63]. T is responsible for differentiation of the Wolffian ducts into internal male reproductive structures in utero (epididymides, vasa deferentia, seminal vesicles, and normal development of the fetal testes) [63], the acquisition of preputial separation, and onset of spermatogenesis in pubertal males. At 11,250 ppm (750 mg/kg per day), all of these endpoints were affected; it is clearly a dose resulting in permanent adverse effects. At 3750 ppm, only AGD was affected, with no effects on the other endpoints; it is not a dose resulting in adverse effects. This study focused on the outcomes/consequences of administering an anti-androgen that interferes with fetal testicular biosynthesis, not on the hormonal or other adjustments (typically transient, at least in adults) made by the animals to these exposures. It must be noted that such alterations in utero may result in permanent changes in the offspring. The high dose in the present study (resulting in BBP intakes of approximately 750 mg/kg per day) was, in fact, selected as the “positive control” (based on the data of Gray et al. [17] and indicated that the test system (CD® (SD) rat) was susceptible, and the laboratory was competent to detect effects if they were present. In addition, the U.S. EPA OPPTS two-generation study design has been criticized since only one F1 male (and one F1 female) per litter is retained to adulthood and necropsied after mating for males (or at weaning of F2 litters for females). Therefore, relatively few animals are evaluated as adults, with the risk that effects present at low incidences and/or visible only in adults could be missed. Since the offspring were initially exposed during gestation and lactation as members of in utero and suckling litters, the appropriate unit for usual statistical comparisons is the litter. The larger the number of litters, the more powerful the analysis becomes. Using only one member/sex/litter, versus more than one member/sex/litter, would increase the possibility that the one member is not representative (or typical) of the litter, but this would increase the variance term of the mean of the litter values and would not affect the statistical power (since the “n” number of litters would not change). In order to increase the statistical power to determine effects, the SUDAAN® software [45] for correlated data was used since it employs the individual animal as the “n” with factors to correct for within-litter clustering (i.e. intralitter correlation, the tendency for littermates to respond similarly). In addition, in this study, culled offspring pups were necropsied on PND 4 (not required by the OPPTS testing guidelines, but very useful). Offspring pups are required to be necropsied at weaning (up to three/sex/litter), and they are required to be examined as adults (minimum one/sex/litter for a total of 30 per sex per group). The number of F1 and F2 offspring

261

examined in toto, including on PND 4 (culls), PND 21 (at weaning), and as adults (F1 only; F2 were terminated at weaning) in this study, involved greater than 250 male and female offspring examined per group for F1 offspring at all doses and for F2 offspring at 0, 750, and 3750 ppm. The reduced number of F2 offspring necropsied at 11,250 ppm (137) was due to reduced numbers of F2 litters and reduced numbers of pups/litter (due to reproductive effects on the F1 parental males). The power of this study design is in the number of offspring examined and in the use of F1 and F2 offspring as “replicates” to confirm effects. The incidences (percentages) of F1 and F2 male reproductive tract malformations by age at evaluation and dose are presented in Fig. 5A (F1) and B (F2). It is obvious that for the BBP-specific malformations, evaluation on PND 4 is not sensitive, although two male F1 pups at 11,250 ppm were detected with such a malformation (ectopic testes), seen rarely to never in controls, on PND 4. Evaluation on PND 21 resulted in statistically significant (and biologically relevant) increases in malformations at 11,250 ppm for both F1 and F2 male offspring. The incidences were higher (and also statistically significant) when offspring were evaluated as adults, obviously due to the ability to better examine the testes, epididymides, and accessory sex organs, and to the possibility that effects progressed (worsened) with continuous exposure over time, but comparable effects were, in fact, observed at the PND 21 necropsy (Fig. 5A and B). For robust assessment of possible reproductive toxicants by many modes of action, with target site(s) central or peripheral, receptor-mediated or not, with exposure continuing over generations and during sensitive life stages, the current U.S. EPA OPPTS [23] two-generation testing guideline is the best study design available. Studies tailored to evaluate specific mechanisms or endpoints are also useful, especially to identify hazard (the intrinsic capacity for the chemical to do harm) and identify sensitive endpoints, but they do not have the breadth, depth, or power that the multigeneration study design does to evaluate the effects with realistic routes, doses, and appropriate timing and duration of exposure (i.e. the risk). The first expert panel of the NTP CERHR was convened in 1999 to evaluate seven phthalates, including BBP (as presented in Section 1). In 2000, the CERHR determined that the data support the designation of BBP [13] as a reproductive toxicant in adult male rats, with a NOAEL of approximately 50 mg/kg per day. However, the CERHR report indicated that a complete evaluation of reproductive and developmental effects for BBP could not be made at that time due to the lack of a multigeneration study [13]. The present study fills that gap.

5. Conclusions Exposure of CD® (SD) rats to dietary BBP for two generations, one litter per generation, at 0, 750, 3750, and

262

R.W. Tyl et al. / Reproductive Toxicology 18 (2004) 241–264

11,250 ppm, resulted in: • F0 and F1 parental systemic toxicity at 11,250 ppm: ◦ Reduced body weights and weight gains; ◦ Increased absolute and relative liver weights; ◦ Histopathologic lesions in livers (consistent with peroxisome proliferation). • F1 (but not F0) parental reproductive toxicity at 11,250 ppm: ◦ Reduced mating and fertility indices in F1 parents to make F2 offspring; ◦ Reduced number of implantations and total and live pups per litter on PND 0 for F2 offspring; ◦ Reduced F1 adult male epididymal sperm motility, progressive motility, sperm concentrations; ◦ Increased incidence of F1 weanling and adult male reproductive tract malformations (from perinatal exposure); ◦ Increased gross and histopathologic lesions in F1 adult male testes, epididymides, and prostate (from perinatal exposure). • F1 and F2 offspring toxicity at 11,250 ppm: ◦ Reduced body weights in F1 and F2 pups during lactation at 11,250 ppm; ◦ Shortened anogenital distance in F1 and F2 males at 11,250 ppm; ◦ Delayed acquisition of puberty in F1 males and females at 11,250 ppm; ◦ Retention of nipples and areolae in preweanling F1 and F2 males at 11,250 ppm; ◦ Male reproductive system malformations at 11,250 ppm. • F1 and F2 offspring effects at 3750 ppm: ◦ Reduced anogenital distance in F1 and F2 males at birth at 3750 ppm; ◦ No other effects at this dose in F1 or F2 males, and no effects in F1 or F2 females at this dose; Therefore, in rats, under the conditions of this study: • The F0 and F1 parental systemic and reproductive NOAEL is 3750 ppm (equivalent to approximately 250 mg/kg per day). • The F1 and F2 offspring reproductive toxicity NOAEL also is 3750 ppm (equivalent to approximately 250 mg/kg per day). • The F1 and F2 offspring reproductive toxicity NOEL (no observable effect level) in males is 750 ppm (equivalent to approximately 50 mg/kg per day), based on the significantly shortened anogenital distance in F1 and F2 male pups at birth at 3750 (and at 11,250) ppm, with no effects on reproductive development, structures, or functions at 3750 mg/kg per day. Acknowledgments The authors wish to thank the staff of RTI’s Reproductive and Developmental Toxicology Group, the Quality As-

surance Unit, and the Materials Handling Facility for their expertise and dedication on this study. This study was sponsored by the European Council for Plasticizers and Intermediates (ECPI), Brussels, Belgium; Dr. David Cadogan, Director. Dr. Tyl’s special thanks go to Ms. Cathee Winkie, Dr. Tyl’s Administrative Coordinator, for her accurate and patient typing (and retyping . . . ) of this manuscript. References [1] Sharpe RM, Fisher JS, Millar MM, Jobling S, Sumpter JP. Gestational and lactational exposure of rats to xenoestrogens results in reduced testicular size and sperm production. Environ Health Perspect 1995;103:1136–43. [2] Jobling S, Reynolds T, White R, Parker MG, Sumpter JP. A variety of environmentally persistent chemicals, including some phthalate plasticizers, are weakly estrogenic. Environ Health Perspect 1995;103:582–7. [3] Harris CA, Henttu P, Parker MG, Sumpter JP. The estrogenic activity of phthalate esters in vitro. Environ Health Perspect 1997;105:802– 11. [4] Koop CE, Juberg DR. Review and consensus statement: a scientific evaluation of the health effects of two plasticizers used in medical devices and toys: a report from the American Council on Science and Health. Medscape Gen Med J 1999;22:1–35. [5] Moore NP. Review: the oestrogenic potential of the phthalate esters. Reprod Toxicol 2000;14(3):183–92. [6] Gray Jr LE, Wolf C, Lambright C, Mann P, Price M, Cooper RL, et al. Administration of potentially antiandrogenic pesticides (procymidone, linuron, iprodione, chlozolinate, p,p -DDE, and ketoconazole) and toxic substances (dibutyl- and diethylhexyl phthalate, PCB 169, and ethane dimethane sulphonate) during sexual differentiation produces diverse profiles of reproductive malformations in the male rat. J Toxicol Ind Health 1999;15(1/2):94– 118. [7] Zacharewski T. Identification and assessment of endocrine disruptors: limitations of in vivo and in vitro assays. Environ Health Perspect 1998;106(2):577–82. [8] Blount BC, Silva MJ, Caudill SP, et al. Levels of seven urinary phthalate metabolites in a human reference population. Environ Health Perspect 2000;108(10):972–82. [9] DHHS (Department of Health and Human Services) report. Second National Report on Human Exposure to Environmental Chemicals, National Center for Environmental Health; 2003. [10] Barr DB, Manori S, Kayoko K, Reidy J, Malek N, Hurtz D, et al. Assessing human exposure to phthalates using monoesters and their oxidized metabolites as biomarkers. Environ Health Perspect 2003; Doi: 10.1289/ehp.6074 (available at http://dx.doi.org/) online 24 February 2003. [11] David RM. Letter: exposure to phthalate esters. Environ Health Perspect 2000;108(10):A-440. [12] Kohn MC, Parham F, Masten SA, Portier CJ, Shelby MD, Brock JW, et al. Letter: human exposure estimates for phthalates. Environ Health Perspect 2000;108(10):A440–1. [13] Kavlock R, Boekelheide K, Chapin R, et al. NTP Center for the Evaluation of Risks to Human Reproduction: phthalates expert panel report on the reproductive and developmental toxicity of butyl benzyl phthalate. Reprod Toxicol 2002;16(5):453–87. [14] Ward JM, Peters JM, Perella CM, Gonzalez FJ. Receptor and nonreceptor-mediated organ-specific toxicity of di(2-ethylhexyl) phthalate (DEHP) in peroxisome proliferator-activated receptor ␣-null mice. Toxicol Pathol 1998;26:240–6. [15] NTP (National Toxicology Program). Developmental toxicity evaluation of butyl benzyl phthalate administered in the feed to CD rats on gestation days 6–15; 1989. NTP Report No. 89-246.

R.W. Tyl et al. / Reproductive Toxicology 18 (2004) 241–264 [16] NTP (National Toxicology Program). Developmental toxicity evaluation of butyl benzyl phthalate in CD-1 Swiss mice; 1990. NTP Report No. 90-114. [17] Piersma AH, Verhoef A, Te Biesebeek JD, Peters MN, Slob W. Developmental toxicity of butyl benzyl phthalate in the rat using a multiple dose study design. Reprod Toxicol 2000;14(5):417–25. [18] Mylchreest E, Cattley RC, Sar M, Foster PM. The effects of di(butyl) phthalate on prenatal androgen-regulated male sexual differentiation are not mediated by direct interaction with the androgen receptor. Teratology 1998;57(4/5):199 (Abstract No. 25).. [19] Mylchreest E, Cattley RC, Foster PM. Male reproductive tract malformations in rats following gestational and lactational exposure to di(n-butyl) phthalate: an antiandrogenic mechanism? Toxicol Sci 1998;43(1):47–60. [20] Mylchreest E, Sar M, Cattley RC, Foster PM. Disruption of androgen-regulated male reproductive development by di(n-butyl) phthalate during late gestation in rats is different from flutamide. Toxicol Appl Pharmacol 1999;156:81–95. [21] Parks LG, Ostby JS, Lambright CR, Abbott BD, Klinefelter GR, Barlow NJ, et al. The plasticizer diethylhexyl phthalate induces malformations by decreasing fetal testosterone synthesis during sexual differentiation in the male rat. Toxicol Sci 2000;58:339–49. [22] Gray Jr LE, Ostby J, Wolf C, Lambright C, Kelce W. Annual review. The value of mechanistic studies in laboratory animals for the production of reproductive effects in wildlife: endocrine effects on mammalian sexual differentiation. Environ Toxicol Chem 1998;17(1):109–18. [23] Gray Jr LE, Ostby J, Price M, Veeramachaneni DNR, Parks L. Perinatal exposure to the phthalates DEHP, BBP and DINP, but not DEP, DMP or DOTP alters sexual differentiation of the male rat. Toxicol Sci 2000;58(2):350–65. [24] Gray Jr LE, Ostby J, Furr J, et al. Review: effects of environmental antiandrogens on reproductive development in experimental animals. Hum Reprod Update 2001;7(3):248–64. [25] Ema M, Miyawaki E, Kawashima K. Critical period for adverse effects on development of di-n-butyl phthalate during late pregnancy. Toxicol Lett 2000;111:271–8. [26] Ema M, Miyawaki E, Kawashima K. Reproductive effects of butyl benzyl phthalate in pregnant and pseudopregnant rats. Reprod Toxicol 1998;12(2):127–32. [27] Ema M, Miyawaki E. Effects on development of the reproductive system in male offspring of rats given butyl benzyl phthalate during late pregnancy. Reprod Toxicol 2002;16(1):71–6. [28] Nagao T, Ohta R, Marumo H, Shindo T, Yoshimura S, Ono H. Effect of butyl benzyl phthalate in Sprague–Dawley rats after gavage administration: a two-generation reproductive study. Repro Toxicol 2000;14(6):513–32. [29] U.S. Environmental Protection Agency (EPA). Office of Prevention, Pesticides and Toxic Substances (OPPTS), Health Effects Test Guidelines, OPPTS 870.3800. Reproduction and fertility effects [Final Guideline, August 1998]. [30] Mylchreest E, Wallace DG, Cattley RC, Foster PMD. Dose-dependent alterations in androgen-regulated male reproductive development in rats exposed to di(n-butyl) phthalate during late gestation. Toxicol Sci 2000;55:143–51. [31] NRC (National Research Council). Guide for the care and use of laboratory animals. institute of laboratory animal resources. Commission on Life Sciences, National Academy Press, National Institutes of Health. Revised 1996. [32] Pederson T, Peters H. Proposal for a classification of oocytes and follicles in the mouse ovary. J Reprod Fertil 1968;17:555–7. [33] Heindel JJ, Thomford PJ, Mattison DR. Histological assessment of ovarian follicle number in mice as a screen for ovarian toxicity. In: Hirshfield AN, editor. Growth factors and the ovary. New York: Plenum Press; 1989. p. 421–6. [34] Zeger S, Liang K. Longitudinal data analysis for discrete and continuous outcomes. Biometrics 1986;42:121–30.

263

[35] Royall RM. Model robust confidence intervals using maximum likelihood estimators. Int Statist Rev 1986;54:221–6. [36] Huber PJ. The behavior of maximum likelihood estimates under nonstandard conditions. In: Proceedings of the Fifth Berkeley Symposium on Mathematical Statistics and Probability, vol. 1; 1967. p. 221–33. [37] Levene H. Robust tests for the equality of variance. In: Olkin I, editor. Contributions to probability and statistics. Palo Alto: Stanford University Press; 1960. p. 278–92. [38] SAS Institute Inc. SAS® language and procedures: usage, version 6, 1st ed. Cary, NC: SAS Institute Inc.; 1989. [39] SAS Institute Inc. SAS/STAT® users’ guide, version 6, 4th ed., vols. 1 and 2. Cary, NC: SAS Institute Inc.; 1989. [40] SAS Institute Inc. SAS® language: reference, version 6, 1st ed. Cary, NC, 1990. [41] SAS Institute Inc. SAS® language: procedures guide, version 6, 3rd ed. Cary, NC: SAS Institute Inc.; 1990. [42] SAS Institute Inc. SAS® companion for the VMSTM environment, version 6, 1st ed. Cary, NC: SAS Institute Inc.; 1990. [43] SAS Institute Inc. SAS® companion for the Microsoft Windows environment. Cary, NC: SAS Institute Inc.; 1996. [44] SAS Institute Inc. SAS/STAT® software: changes and enhancements through Release 6.12. Cary, NC: SAS Institute Inc.; 1997. [45] Shah BV, Barnwell BG, Bieler GS. SUDAAN® software for the statistical analysis of correlated data. User’s manual, Release 7.5, vol. 1. Research Triangle Park, NC: Research Triangle Institute; 1997. [46] Dunnett CW. A multiple comparison procedure for comparing several treatments with a control. J Am Stat Assoc 1955;50:1096–121. [47] Dunnett CW. New tables for multiple comparisons with a control. Biometrics 1964;20:482–91. [48] Snedecor GW, Cochran WG. Statistical methods, 6th ed. Ames, IA: Iowa State University Press; 1967. [49] Agresti A. Categorical data analysis. New York: Wiley; 1990. [50] Armitage P. Test for linear trends in proportions and frequencies. Biometrics 1955;11:375–86. [51] Cochran W. Some methods for strengthening the common χ2 tests. Biometrics 1954;10:417–51. [52] Barber ED, Astill BD, Moran EJ, Schneider BF, Gray TJB, Lake BG, et al. Peroxisome induction of seven phthalate esters. Toxicol Ind Health 1987;2:7–22. [53] Woodward KN. Subchronic toxicity. In: Phthalate esters: toxicity and metabolism, vol. I. Boca Raton, FL: CRC Press; 1988. p. 93–122 [chapter 4]. [54] Conney AH. Pharmacological implications of microsomal enzyme induction. Pharmacol Rev 1967;19:317–66. [55] Gallavan Jr RH, Holson JF, Stump DG, Knapp JF, Reynolds VL. Interpreting the toxicological significance of alterations in anogenital distance: potential for confounding effects of progeny body weight. Reprod Toxicol 1999;13:383–90. [56] Quigley CA, DeBellis A, Marschke KB, El-Awady MK, Wilson EA, French FS. Androgen receptor defects: historical, clinical and molecular perspectives. Endocr Rev 1995;16(3):271–321. [57] Gilbert SF. Sex determination. In: Developmental biology, 5th ed. MA: Sinauer Associates, Inc., Publishers; 1997. p. 773–804 [chapter 20]. [58] McIntyre BS, Barlow NJ, Foster PMD. Androgen-mediated development in male rat offspring exposed to flutamide in utero: permanence and correlation of early changes in anogenital distance and nipple retention with malformations in androgen-dependent tissues. Toxicol Sci 2001;62(2):236–49. [59] Gray Jr LE, Kelce WR, Wiese T, et al. Endocrine screening methods workshop report: detection of estrogenic and androgenic hormonal and antihormonal activity for chemicals that act via receptor or steroidogenic enzyme mechanisms. Reprod Toxicol 1997;11:719–50. [60] Parks LG, Ostby JS, Lambright CR, Abbott BD, Gray Jr LE. Perinatal butyl benzyl phthalate (BBP) and bis(2-ethylhexyl) phthalate (DEHP) exposures induce antiandrogenic effects in Sprague–Dawley (SD) rats. Biol Reprod 1999;60:153.

264

R.W. Tyl et al. / Reproductive Toxicology 18 (2004) 241–264

[61] McIntyre BS, Barlow NJ, Foster PMD. Male rats exposed to linuron in utero exhibit permanent changes in anogenital distance, nipple retention, and epididymal malformations that result in subsequent testicular atrophy. Toxicol Sci 2002;65(1):62–70. [62] Bowman CJ, Barlow NJ, Turner KJ, Wallace DG, Foster PMD. Effects of in utero exposure to finasteride on androgen-dependent reproductive development in the male rat. Toxicol Sci 2003;74(2): 393–406. [63] U.S. Environmental Protection Agency (EPA). Part II: Environmental protection agency: reproductive toxicity risk assessment guidelines; notice. Federal Register 1996;61(212):56274–322. [64] Tyl RW, Myers CB, Marr MC, Brine DR, Fail PA, Seely JC, et al. Two-generation reproduction study with para-tert-octylphenol in rats. Regul Toxicol Pharmacol 1999;30:81–95. [65] Tyl RW, Myers CB, Marr MC, et al. Three-generation reproductive toxicity study of dietary bisphenol A (BPA) in CD® (Sprague–Dawley) rats. Toxicol Sci 2002;68(1):121–46. [66] Kennedy GC, Mitra J. Body weight and food intake as initiating factors for puberty in the rat. Physiology 1963;166:408–18. [67] Carney EW, Scortichini BS, Crissman JW. Feed restriction during in utero and neonatal life: effects on reproductive and developmental end points in the CD rat. Toxicologist 1998;42:102–3 [Abstract no. 506]. [68] Shultz VP, Phillips SL, Foster PMD, Gaido KW. Microarray analysis of altered gene expression in the testes of fetal rats exposed to DBP. Teratology 2000;61(6):455 [Abstract no. 57]. [69] Shultz VP, Phillips SL, Sar M, Foster PMD, Gaido KW. Altered gene profiles in fetal rat testes after in utero exposure to di(n-butyl) phthalate. Toxicol Sci 2001;64(2):233–42. [70] Tarka DK, Klinefelter GR, Suarez J, Rogers JM. Effect of developmental exposure to ethane dimethane sulphonate (EDS), bromochloroacetic acid (BCA) and molinate in whole body testosterone levels in CD-1 mice. Teratology 2000;61(6):437 [Abstract no. 5]. [71] Li LH, Jester WF, Laslett AL, Orth JM. A single dose for di-(2-ethylhexyl) phthalate in neonatal rats alters gonocytes, reduces Sertoli cell proliferation, and decreases cyclin D2 expression. Toxicol Appl Pharmacol 2000;166(3):222–6. [72] Imajima T, Shono T, Zakaria O, Suita P. Prenatal phthalate causes cryptorchidism postnatally by inducing transabdominal ascent of the testis in fetal rats. J Pediatr Surg 1997;32:18–21.

[73] Shono T, Kai H, Sulta H, Nawata S. Time-specific effects of mono-n-butyl phthalate on the transabdominal descent of the testis in rat fetuses. BJU Int 2000;86:121–5. [74] Ema M, Miyawaki E, Hirose A, Kamata E. Decreased anogenital distance and increased incidence of undescended testes in fetuses of rats given monobenzyl phthalate, a major metabolite of butyl benzyl phthalate. Reprod Toxicol 2003;17(4):407–12. [75] Kubota Y, Temelcos C, Bathgate RA, Smith KJ, Scott D, Zhao C, et al. The role of insulin 3, testosterone, Müllerian inhibiting substance and relaxin in rat gubernacular growth. Mol Hum Reprod 2002;8(10):900–5. [76] Nef S, Parada LF. Cryptorchidism in mice mutant for Insl3. Nat Genet 1999;22(3):295–9. [77] Zimmermann S, Steding G, Emmen JM, Brinkman AO, Nayernia K, Holstein AF, et al. Targeted disruption of the Insl3 gene causes bilateral crytorchidism. Mol Endocrinol 1999;13(5):681–91. [78] Emmen JM, McLuskey A, Adham IM, Engel W, Verhoef-Post M, Themmen AP, et al. Involvement of insultin-like factor 3 (Insl3) in diethylstilbestrol-induced cryptorchidism. Endocrinology 2000;141(2):846–9. [79] Tomboc M, Lee PA, Mitwally MF, Schneck FX, Bellinger M, Witchel SF. Insulin-like 3/relaxin-like factor gene mutations are associated with cryptorchidism. J Clin Endocrinol Metab 2000;85(11):4013–8. [80] Baker LA, Nef S, Nguyen MT, Stapleton R, Pohl H, Parada LF. The insulin-3 gene: lack of a genetic basis for human cryptochidism. J Urol 2002;167(6):2534–7. [81] Emmen JM, McLuskey A, Adham IM, Engel W, Grootegoed JA, Brinkman AO. Hormonal control of gubernaculum development during testis descent: gubernaculum outgrowth in vitro requires both insulin-like factor and androgen. Endocrine 2000;14(12):4720–7. [82] Gray Jr LE, Foster P. Significance of experimental studies for assessing adverse effects of endocrine disrupting chemicals. In: SCOPE/IUPAC International Symposium on Endocrine Auto Substances and Supplementary Workshop, Yokohama, Japan. Program and Collection Abstract, Subtopic 2002;14:40–1. [83] Wilson VS, Lambright C, Furr J, Ostby J, Wood C, Held G, et al. Phthalate ester-induced gubernacular ligament lesions are associated with reduced Insl3 gene expression in the fetal rat testis during sexual differentiation. Tox Lett 2004;146:207–15. [84] Ema M, Itami T, Kawasaki H. Teratogenic evaluation of butyl benzyl phthalate in rats. Toxicol Lett 1992;61:1–7.