Hazard to the developing male reproductive system from cumulative exposure to phthalate esters—dibutyl phthalate, diisobutyl phthalate, butylbenzyl phthalate, diethylhexyl phthalate, dipentyl phthalate, and diisononyl phthalate

Hazard to the developing male reproductive system from cumulative exposure to phthalate esters—dibutyl phthalate, diisobutyl phthalate, butylbenzyl phthalate, diethylhexyl phthalate, dipentyl phthalate, and diisononyl phthalate

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Regulatory Toxicology and Pharmacology 53 (2009) 90–101

Contents lists available at ScienceDirect

Regulatory Toxicology and Pharmacology journal homepage: www.elsevier.com/locate/yrtph

Hazard to the developing male reproductive system from cumulative exposure to phthalate esters—dibutyl phthalate, diisobutyl phthalate, butylbenzyl phthalate, diethylhexyl phthalate, dipentyl phthalate, and diisononyl phthalate q Robert Benson * Evergreen, CO 80439, USA

a r t i c l e

i n f o

Article history: Received 5 September 2008 Available online 11 December 2008 Keywords: Dibutyl phthalate Diisobutyl phthalate Butylbenzyl phthalate Diethylhexyl phthalate Dipentyl phthalate Diisononyl phthalate Male reproductive and developmental effects

a b s t r a c t Phthalate esters are found in a wide variety of consumer and food packing products. Hence there is widespread exposure of the human population to these chemicals. Some of the phthalate esters are known to be toxic to the developing male reproductive system. This paper derives a reference dose (RfD) for each of the phthalate esters (dibutyl phthalate, diisobutyl phthalate, butylbenzyl phthalate, diethylhexyl phthalate, dipentyl phthalate, and diisononyl phthalate) that cause these effects. As these phthalate esters cause similar adverse biological effects and have the same mechanism of action, it is appropriate in a risk assessment to consider the potential adverse effects from cumulative exposure to these chemicals using a dose addition model. This paper provides examples of a cumulative risk assessment using the hazard index and relative potency approaches from the RfDs derived from studies in laboratory animals and exposure information in people. The results of the cumulative risk assessments for both a US and a German population show that the hazard index is below one. Thus it is unlikely that humans are suffering adverse developmental effects from current environmental exposure to these phthalate esters. Published by Elsevier Inc.

1. Introduction Phthalate esters are used as plasticizers to impart flexibility and resilience to plastic products. Many consumer products and food packaging products contain phthalate esters. The major uses of selected phthalate esters are summarized in Table 1. Information on the use of dipentyl phthalate was not located. The phthalate esters are not covalently bound in the products in which they are used and can leach into the surrounding environment. Because the phthalate esters are used in such a wide variety of consumer products, human exposure to the phthalate esters is widespread. The phthalate diesters are rapidly converted to the monoesters in the rat (Rowland et al., 1977) and in the human (Anderson et al., 2001). For those phthalate esters that cause developmental effects, the monoester metabolite is believed to be the toxic chemical

q The author is employed at a Federal Regulatory Agency. However, the views expressed in this paper are those of the author and do not necessarily reflect the views and policies of that agency. Mention of trade names or commercial products does not constitute endorsement or recommendation for use. This project received no outside source of funds. The author declares he has no competing financial interests. * Fax: +1 303 312 6116. E-mail address: [email protected]

0273-2300/$ - see front matter Published by Elsevier Inc. doi:10.1016/j.yrtph.2008.11.005

(Foster, 2006). Measurement of the urinary concentration of the monoester metabolites and additional oxidation products has been included in the National Health and Nutritional Examination Survey (NHANES). Data are available from a subset of samples collected in 1988–1994 (Blount et al., 2000) and from nationwide sampling conducted in 1999–2000 and 2001–2002 (DHHS, CDC, 2005). Exposure data are also available for a German population (Wittassek and Angerer, 2008). US EPA’s Integrated Risk Information System (IRIS) website often provides a summary of toxicological data on a chemical and a numerical estimate of toxicity that can be used in a risk assessment. Unfortunately, the assessments for the phthalate esters have not been updated since the late 1980s. The posted reference dose (RfD) for dibutyl phthalate (0.1 mg/kg-day) is based on a NOAEL (No Observed Adverse Effect Level) of 125 mg/kg-day and LOAEL (Lowest Observed Adverse Effect Level) of 600 mg/kg-day for mortality in rats and incorporates a total uncertainty factor of 1000. The posted RfD for butylbenzyl phthalate (0.2 mg/kg-day) is based on a NOAEL of 159 mg/kg-day and a LOAEL of 470 mg/kg-day for increased liver weight in rats and incorporates a total uncertainty factor of 1000. The posted RfD for diethylhexyl phthalate (0.02 mg/ kg-day) is based on a LOAEL of 19 mg/kg-day for increased liver weight in guinea pigs and incorporates a total uncertainty factor of 1000. Since these files were posted on the IRIS site, a large number of studies have been published showing that dibutyl phthalate,

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Table 1 Phthalate diester, its corresponding monoester metabolite, and uses*. Phthalate diester

Monoester metabolite

Major uses

Dibutyl phthalate (DBP) CAS # 84-74-2 Mol wt. 278

Monobutyl phthalate (MBP)

Diisobutyl phthalate (DiBP) CAS #84-695 Mol wt. 278 Butylbenzyl phthalate (BBP) CAS # 85-687 Mol wt. 312 Diethylhexyl phthalate (DEHP) CAS # 117-81-7 Mol wt. 390

Monoisobutyl phthalate (MiBP)

Plastic products containing nitrocellulose, polyvinyl acetate, or polyvinyl chloride; pharmaceutical coatings; lubricant for aerosol valves; anti-forming agent; skin emollient; nail polishes; fingernail elongators; and hair spray Same as dibutyl phthalate but less commonly

Diisononyl phthalate (DiNP) CAS # 68515-48-0 and 28553-12-0 Mol wt. 419 *

Monobutyl phthalate (MBP) and monobenzyl phthalate (MBzP) Monoethylhexyl phthalate (MEHP)

Monoisononyl phthalate (MiNP)

Polyvinyl chloride products including vinyl tile, food conveyer belts, carpet tile, artificial leather, tarps, automotive trim, weather strippers, traffic cones, and vinyl gloves Polyvinyl chloride products including building products, car products, clothing, food packaging, children’s products, and in medical devices (storage containers, bags, and flexible tubing) Flexible polyvinyl chloride products including children’s toys, flooring, gloves, food packaging material, drinking straws, and garden hoses

Comparable information was not found for dipentyl phthalate.

butylbenzyl phthalate, and diethylhexyl phthalate cause effects on the developing male reproductive system in laboratory animals at exposures below those causing systemic effects. EPA has provided the external review draft and peer review report for the Toxicological review of dibutyl phthalate ( US EPA, 2006). However, the final assessment has not been posted on the IRIS site. Although butylbenzyl phthalate and diethylhexyl phthalate have been listed on the IRIS agenda for many years, no documents are available for public review. There is no IRIS file for diisobutyl phthalate, dipentyl phthalate, or diisononyl phthalate and neither chemical appears on the IRIS agenda. Matsumoto et al. (2008) summarized the studies available in human populations on exposure to phthalate esters and health effects. A number of studies show a correlation between effects (sperm and semen parameters in adults; anogenital distance in newborn males; and serum hormone levels in newborn infants) and environmental exposure to phthalate esters. Many of these same endpoints are observed in controlled studies in laboratory animals. However, in some cases, correlations with the same endpoints and exposure to phthalate esters were observed with phthalate esters (e.g., monomethyl phthalate and monoethyl phthalate) that do not cause these effects in laboratory animals. After a thorough review of these studies, Matsumoto et al. (2008) concluded: ‘‘some of the findings in human populations are consistent with animal data suggesting that PAEs (phthalic acid esters) and their metabolites produce toxic effects in the reproductive system. However, it is not yet possible to conclude whether phthalate exposure is harmful for human reproduction.” A review of these studies in human populations is not included in this paper. However, based on the data available from laboratory animals, there is concern that environmental exposure to phthalate esters could cause adverse effects on the developing human male reproductive system. Additional concern is raised by the demonstration of Silva et al. (2004) that phthalate monoesters known to cause developmental toxicity in rats are detected in a large percentage of human amniotic fluid samples. A literature review was conducted for the phthalate esters that have the potential to be developmental toxicants based on structure-activity relationships. The focus was on phthalate esters with a side chain containing four to ten carbons in the ortho position (Foster et al., 1980; Gray et al., 2000). The shorter chain phthalate esters (dimethyl phthalate and diethyl phthalate) were not included because these chemicals are not developmental toxicants. Relevant toxicological data following in utero exposure was located only for dibutyl phthalate, diisobutyl phthalate, butylbenzyl phthalate, diethylhexyl phthalate, dipentyl phthalate, and diisononyl phthalate. This paper does not contain a review of all of the studies available on these chemicals. The focus is on the most

sensitive effect in the most sensitive life stage—the reproductive tract of the developing male fetus—to allow development of the RfD based on the same toxicological endpoint in the same life stage. A major objective of this paper is to provide information on the hazard identification, dose–response using either the NOAEL/ LOAEL or benchmark dose approaches, and to derive a RfD for effects on the developing male reproductive system following in utero exposure of laboratory animals to dibutyl phthalate, diisobutyl phthalate, butylbenzyl phthalate, diethylhexyl phthalate, dipentyl phthalate, and diisononyl phthalate. The development of the male reproductive tract is dependent on the presence of testosterone and the androgen receptor (reviewed in Foster, 2006; Hughes, 2001; Hughes et al., 2001). Any chemical that reduces the concentration of the androgen receptor–testosterone complex during the critical developmental window has the potential of causing malformations of the male reproductive tract. The mechanism of action for the adverse developmental effects for each of the phthalate esters included in this paper is the decrease in the concentration of fetal testosterone during the critical developmental window (Foster, 2006). There is experimental evidence that the decrease in fetal testosterone caused by cumulative exposure to these six phthalate esters in rats follows a dose addition model (Howdeshell et al., 2007; 2008). Therefore, it is appropriate to conduct a cumulative risk assessment for simultaneous exposure to these phthalate esters using a dose addition model and the RfDs derived in this paper. Accordingly, a second major objective of this paper is to provide examples of a cumulative risk assessment using exposure information from a US and a German population. 2. Methods 2.1. Calculation of the reference dose (RfD) EPA defines the RfD as ‘‘an estimate (with uncertainty spanning perhaps an order of magnitude) of a daily oral exposure to the human population (including sensitive subgroups) that is likely to be without an appreciable risk of deleterious effects during a lifetime. It can be derived from a NOAEL, LOAEL, or benchmark dose, with uncertainty factors generally applied to reflect limitations of the data used” (US EPA, 2008a). The RfD is calculated by dividing the NOAEL, LOAEL, or benchmark dose by uncertainty factors. EPA defines an uncertainty factor as ‘‘one of several, generally 10-fold, default factors used in operationally deriving the RfD and RfC from experimental data.” For the chemicals included in this paper, the uncertainty factors include a factor of 10 for intraspecies variability, a factor of 10 for

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interspecies variability, and, if needed, a factor of 10 for extrapolation from a LOAEL to a NOAEL. There is not sufficient robust information to reduce the interspecies or intraspecies uncertainty factors from the default value of 10. Information required to reduce the pharmacokinetic portion of the uncertainty factors would include quantitative information in laboratory animals and humans on adsorption, metabolism (including glucuronidation), distribution, and excretion. It would also be necessary to have information on the variability in pharmacokinetics across the human population. Information required to reduce the pharmacodynamic portion of the uncertainty factors would include quantitative information in laboratory animals and humans showing that the same degree of change in testosterone in the fetal testes would give the same incidence of biological changes in each species. It would also be necessary to have information on the variability of testosterone in the fetal testes in humans. For the chemicals included in this paper, the most vulnerable target for adverse health effects is the reproductive tract of the developing male fetus. Therefore, the RfD is based on exposure-response data from studies following in utero exposure. This RfD is protective for systemic effects that occur at a higher exposure during neonatal or adult life. 2.2. Dose–response evaluation The NOAEL and LOAEL were assigned based on the exposures used in the laboratory animal study. If appropriate, benchmark dose analysis was conducted using EPA software (US EPA, version 1.4.1c) (US EPA, 2008b). The mathematic models included in the software are gamma, logistic, log-logistic, multi-stage, probit, log-probit, quantal-linear, and Weibull for dichotomous data and linear, polynomial, power, and Hill for continuous data. The choice of the best fitting model is based on p value (p > 0.1), visual fit, and lowest AIC (Akaike’s Information Criterion) (US EPA, 2000). 2.3. Calculation of relative potency factor The RfD for diethylhexyl phthalate is assigned a relative potency factor of 1. Values for the other phthalate esters are calculated by dividing the RfD for diethylhexyl phthalate in mmol/kgday by the RfD for the other phthalate ester in mmol/kg-day. The values are rounded to two significant digits. 2.4. Calculation of diethylhexyl phthalate equivalents Diethylhexyl phthalate equivalents are calculated by dividing the estimated exposure to the phthalate ester in mg/kg-day by the molecular weight of the phthalate ester and then multiplying by the relative potency factor.

3. Dibutyl phthalate—hazard identification, dose–response, and RfD The National Toxicology Program Center for the Evaluation of Risk to Human Reproduction (NTP-CERHR) has provided a summary of the toxicological data on dibutyl phthalate (NTP-CERHR, 2003a; Kavlock et al., 2002a). US EPA (2006) has also summarized these data and included additional data published through February 2006. In the developing male fetus, the toxicological effects observed include a variety of malformations of the male reproductive system in structures that are dependent on properly functioning Leydig cells and the presence of adequate levels of testosterone. The malformations commonly observed in rats include hypospadias; decrease in anogenital distance; delayed preputial separation; agenesis of the prostate, epididymis, and vas deferens; degeneration of the seminiferous epithelium; interstitial cell hyperplasia of the testis; and retention of thoracic areolas or nipples. The underlying mode of action for these effects is inhibition of cholesterol transport and testosterone synthesis in the fetal Leydig cells (Thompson et al., 2005; 2004; Lehmann et al., 2004). Lehmann et al. (2004) investigated the exposure-response relationships for the effect of dibutyl phthalate on steroidogenesis in fetal rat testes. Pregnant Sprague-Dawley rats (n = 7 in control and n = 5 in each exposed group) were treated with dibutyl phthalate by gavage in corn oil at 0, 0.1, 1.0, 10, 50, 100, or 500 mg/kg-day on GDs 12–19. The day sperm was detected in the vaginal smear is defined as GD 0. This phase of the study measured gene expression and protein synthesis. Separate groups of animals (n = 7) were treated with dibutyl phthalate by gavage in corn oil at 0, 0.1, 1.0, 10, 30, 50, 100, or 500 mg/kg-day on GDs 12–19 to determine fetal testosterone concentration. Testes were isolated on GD 19. Changes in gene and protein expression on GD 19 were quantified by RT-PCR and Western blot analysis. Fetal testicular testosterone concentration on GD 19 was determined in three to four individual fetuses from one to four litters per exposure group with a radioimmunoassay after the testes were homogenized and extracted with ethyl acetate and chloroform (4:1). After drying, the extract was dissolved in methanol. The data (Table 2) are reported as the concentration of testosterone in the final methanol solution. Exposure to dibutyl phthalate at 50 mg/kg-day and above resulted in significant reductions in mRNA and protein concentration for proteins and enzymes involved in cholesterol transport and synthesis of testosterone, including scavenger receptor class B1 (SR-B1), steroidogenic acute regulated protein (StAR), P450 side chain cleavage enzyme (P450scc) and 3b-Hydroxysteroid dehydrogenase (3b-HSD). As shown in Table 2, there was a statistically significant decrease (p < 0.05) in mean testosterone concentration at 50 mg/kg-day but not at 30 mg/kg-day. The NOAEL in this study is 30 mg/kg-day and the LOAEL is 50 mg/kg-day. The data from Lehmann et al. (2004) on the concentration of testosterone in the fetal

2.5. Calculation of hazard index using relative potency This hazard index is calculated by dividing the total exposure to diethylhexyl phthalate equivalents by the RfD for diethylhexyl phthalate. 2.6. Hazard quotient (HQ) The HQ for each chemical is the estimated human exposure divided by the RfD and rounded to one significant digit. 2.7. Hazard index (HI) The HI is the sum of the HQs for the individual chemicals and rounded to one significant digit.

Table 2 Testosterone concentration in fetal testis at GD 19 (Lehmann et al., 2004). Exposure to DBP (mg/kg-day)

Testosterone concentration in methanol extract (ng/mL)

SEM

0 0.1 1 10 30 50 100 500

15.8 17.4 15.8 15.5 11.7 6.2* 4.9* 1.1*

1.55 1.41 1.38 1.55 1.62 2.00 1.51 1.55

Source: Lehmann et al. (2004), figure 4, and Hamner Institute Archivist. SEM is standard error of the mean. Statistically different from control (p < 0.05).

*

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testis were collected in too few animals to provide reliable benchmark dose analyses. A more recent study (Howdeshell et al., 2008) measured testosterone production following administration of dibutyl phthalate. Pregnant Sprague-Dawley rats were treated from GD 8 to 18 by gavage with 0 (n = 3), 33 (n = 4), 50 (n = 4), 100 (n = 4), 300 (n = 4), or 600 (n = 4) mg/kg-day. The day of a sperm plug positive is defined as GD 1. Ex vivo testosterone production in male fetuses from the litter was measured by radioimmunoassay at GD 18 using a three hour incubation period as described by Wilson et al. (2004). Data were reported as litter mean and standard error. There was no effect on testosterone production at 100 mg/kg-day and below and a statistically significant decrease (p < 0.01) in testosterone production at 300 mg/kg-day and above. Accordingly, the NOAEL is 100 mg/kg-day and the LOAEL is 300 mg/kg-day for this study. I conducted a benchmark dose analysis of the data. The linear and power models gave the best fit with a BMD1SD of 139 mg/kg-day and a BMDL1SD of 104 mg/kg-day. Mylchreest et al. (2000) gave pregnant Sprague-Dawley CD rats dibutyl phthalate by gavage in corn oil at 0, 0.5, 5, 50, or 100 mg/ kg-day (n = 19–20 per group) or 500 mg/kg-day (n = 11) on GDs 12–21. The day sperm was detected in the vaginal smear is defined as GD 0. In male offspring AGD was decreased at 500 mg/kg-day. A statistically significant increase (p < 0.05, using a nested analysis) in retained areolas or nipples on PND 14 was present in 31% and 90% of male pups at 100 and 500 mg/kg-day, respectively (80% and 100% of litters affected, respectively). The data are presented in Table 3. The NOAEL is 50 mg/kg-day and the LOAEL is 100 mg/ kg-day for this study. US EPA (2006) calculated a BMDL10 (benchmark dose at a 10% response rate with 95% confidence) of 40 mg/ kg-day (rounded from 39.6 mg/kg-day) from the data of Mylchreest et al. (2000). Another study (Lee et al., 2004) reported effects at lower exposures. Pregnant Sprague-Dawley rats were administered dibutyl phthalate at dietary concentrations of 0, 20, 200, 2000, or 10000 ppm from GD 15 to PND 21. The authors of the study measured food consumption and the body weight of the dams at GD 15– 20, PND 2–10, and PND 10–21. They reported the range in the exposure to the dams during these periods as 0, 1.5–3.0, 14.4– 28.5, 148.2–290.9, or 712.3–1371.8 mg/kg-day. The day when sperm vaginal plugs were observed is defined as GD 0. Lee et al.

(2004) reported a reduction in spermatocyte development in male pups at PND 21 at all exposures (judged minimal to slight at the two lower exposures). At postnatal week 11, there was loss of germ cell development only at the two highest exposures. At postnatal week 20, there was no statistically significant loss of germ cell development at any exposure (highest exposure not tested). Because the minimal to slight reduction in spermatocyte development at PND 21 in immature males at the two lower exposures did not progress to loss of germ cells at postnatal week 11 in mature males, I do not consider the effect reported at the two lower exposures in immature males at PND 21 to be biologically significant. Accordingly, the NOAEL and LOAEL for loss of germ cell development in this study are 14.4–28.5 mg/kg-day and 148.2–290.9, respectively. Lee et al. (2004) also reported finding vacuolar degeneration of alveolar cells or alveolar atrophy in the mammary gland of male pups at postnatal week 11 at all exposures. At postnatal week 20 there was vacuolar degeneration of alveolar cells or alveolar atrophy in the mammary gland of male pups only at 200 and 2000 ppm. Animals exposed at 10000 ppm were not examined. As the biological relevance of the histopathological changes reported in the male mammary gland is unclear, these effects are not considered adverse in this study. The studies considered for the derivation of the RfD for dibutyl phthalate are summarized in Table 4. The RfD for dibutyl phthalate is derived from Lehmann et al. (2004) and the decrease in fetal testosterone with a NOAEL of 30 mg/kg-day and a LOAEL of 50 mg/kgday. While the biochemical change in testosterone concentration is a precursor event, preventing the decrease in testosterone in the developing fetus will prevent the development of gross malformations that will occur at higher exposure. Although it is not known with certainty the degree of testosterone decrease that will trigger the cascade of adverse structural changes in the male fetus after exposure to dibutyl phthalate, it is not unreasonable to assign 50 mg/kg-day as the LOAEL as this is the highest NOAEL for retained areolas or nipples in Mylchreest et al. (2000). A total uncertainty factor of 100 (10 each for interspecies and intraspecies extrapolation) will be used. Accordingly, the RfD for dibutyl phthalate is 0.3 mg/kg-day (equivalent to 0.00108 mmol/kg-day using a molecular weight of 278). 4. Diisobutyl phthalate—hazard identification, dose–response, and RfD

Table 3 Incidence of retained areolas or nipples in male pups at PND 14 (Mylchreest et al., 2000). Exposure to DBP (mg/kg-day)

Pups affected/total pups

Litters affected/total litters

0 0.5 5 50 100 500

9/134 8/119 13/103 12/120 44/141* 52/58*

5/19 5/20 8/19 10/20 16/20* 11/11*

Source: Adapted from Mylchreest et al. (2000), Figure 2. * Statistically different from control (p < 0.05).

Borch et al. (2006a) established that diisobutyl phthalate causes adverse effects on the developing male reproductive system similar to those caused by dibutyl phthalate. Borch et al. (2006a) treated four groups of pregnant Wistar rats by oral gavage in corn oil at 0 or 600 mg/kg-day from GD 7 to either GD 19 or 20/21. The day following mating is defined as GD 1. Male offspring were examined on GD 19 or 20/21 for effects on testosterone production, testicular histopathology, and anogenital distance. Testicular testosterone content and ex vivo production were statistically significantly reduced at GD 20/21, but not at GD 19. Testicular testosterone content was reduced to 9% of the control value and testicular testosterone production was reduced to 4% of the control value.

Table 4 Studies considered for the derivation of the reference dose for dibutyl phthalate. Critical Effect

NOAEL mg/kg-day

LOAEL mg/kg-day

BMDL mg/kg-day

Reference

Decreased fetal testosterone Increase in retained areolas/nipples in males Loss of germ cell development at PNW 11 Decreased fetal testosterone production

30 50 14.4–28.5 100

50 100 148–290.9 300

Not calculated 40 Not calculated 104

Lehmann et al. (2004) Mylchreest et al. (2000) Lee et al. (2004) Howdeshell et al. (2008)

Lehmann et al. (2004) is chosen over the other studies because it has the lowest NOAEL and LOAEL.

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There was a statistically significant increase in testicular histopathology (clustering of small Leydig cells at GD 19 and GD 20/21; vacuolization of Sertoli cells at GD 20/21; and presence of multinuclear gonocytes at GD 20/21). The incidence of each of these effects approached 100% at GD 20/21 after exposure to diisobutyl phthalate at 600 mg/kg-day. There was also a statistically significant reduction in immunohistochemical staining for steroidogenic acute regulated protein (StAR) in Leydig cells at GD 19 and GD 20/21 and a statistically significant reduction in immunohistochemical staining for P450 side chain cleavage enzyme (P450scc) in Leydig cells at GD 20/21. StAR and P450scc are proteins that function in the rate limiting steps in testosterone production in the Leydig cells. There was also a statistically significant reduction in anogenital distance in male pups (normalized for differences in body weight by expressing the results as the anogenital distance divided by the cube root of the body weight) at GD 20/21. There was also a statistically significant increase in the anogenital distance in female pups at GD 19 and GD 20/21. The LOAEL in this study is 600 mg/kg-day, the only exposure studied. A more recent study (Howdeshell et al., 2008) measured testosterone production following administration of diisobutyl phthalate. Pregnant Sprague-Dawley rats were treated from GD 8 to 18 by gavage with 0 (n = 5), 100 (n = 8), 300 (n = 8), 600 (n = 5), or 900 (n = 5) mg/kg-day. The day of a sperm plug positive is defined as GD 1. Ex vivo testosterone production in male fetuses from the litter was measured by radioimmunoassay at GD 18 using a 3 h incubation period as described by Wilson et al. (2004). Data were reported as litter mean and standard error. There was no effect on testosterone production at 100 mg/kg-day and a statistically significant decrease (p < 0.01) in testosterone production at 300 mg/kg-day and above. The NOAEL is 100 mg/kg-day and the LOAEL is 300 mg/kg-day in this study. I conducted a benchmark dose analysis of the data. The Hill model gave the best fit with a BMD1SD of 136 mg/kg-day and a BMDL1SD of 80 mg/kg-day. The studies considered for the derivation of the RfD for diisobutyl phthalate are summarized in Table 5. Borch et al. (2006a) is limited because only one exposure level was used. Howdeshell et al. (2008) provided full exposure-response data and is used to derive the RfD. Using the BMDL1SD of 80 mg/kg-day and a total uncertainty factor of 100 (10 each for interspecies and intraspecies extrapolation), the RfD is 0.8 mg/kg-day, equivalent to 0.00288 mmol/kg-day using a molecular weight of 278. 5. Butylbenzyl phthalate—hazard identification, dose–response, and RfD NTP-CERHR has provided a summary of the toxicological data on butylbenzyl phthalate (NTP-CERHR, 2003b; Kavlock et al., 2002b). The expert panel concluded that there were sufficient data to conclude that butylbenzyl phthalate can cause developmental toxicity in rats and mice, and reproductive toxicity in mice. However, there were inadequate data to determine the potential hazard from perinatal exposure. The panel identified a multigeneration study as a critical data need for butylbenzyl phthalate.

Since the review by the expert panel, two research groups (Nagao et al., 2000; Tyl et al., 2004) have reported the results from two-generation reproduction studies. These two studies document effects on the male reproductive system following exposure during the late gestational period, the critical window for the development of the male reproductive system in rodents. Nagao et al. (2000) conducted a two-generation reproductive toxicity study in Sprague-Dawley rats (25 per sex per dose). Butylbenzyl phthalate was administered by gavage in corn oil at 0, 20, 100, or 500 mg/kg-day. In the parent animals (F0), there was a decrease in relative ovary weight in females at 500 mg/kg-day, but no macroscopic or microscopic changes were found in the reproductive system of males. There was no effect on estrous cyclicity, fertility, or lactation in F0. In the first generation (F1), there was a decrease in mean pup weight at PND 0 for both females and males at 100 and 500 mg/kg-day. This effect was not observed in the F2. In the F1 the anogenital distance at birth was decreased in male pups to 92.3% of control at 500 mg/kg-day. Preputial separation was delayed in males at 500 mg/kg-day from 43.2 ± 1.5 days to 44.5 ± 2.3 days. Vaginal opening for females was not affected at any exposure. Butylbenzyl phthalate did not affect reproductive ability at any exposure in F1 adults. In adult F1 males exposed to 500 mg/kg-day, there were macroscopic and microscopic abnormalities of the testis (atrophy of the seminiferous tubule, decrease in germ cells in the seminiferous tubule, and edema of the interstitium); a decrease in sperm in the epididymis; and a decrease in serum testosterone. The authors concluded that 20 mg/kg-day was the NOAEL in this study. However, the only effects noted at 100 mg/kg-day, the next higher exposure, were an increase in relative kidney weight in F0 females by 8% with a lesser increase at 500 mg/kg-day; an increase in FSH in F0 males by 22% with no further increase at 500 mg/kg-day; a decrease in body weight by 7% and increase in relative kidney weight in F1 adult males by 9%; and a decrease in mean pup weight in F1 at PND 0 by 6% with no further decrease at 500 mg/kg-day. All of these effects are of questionable biological significance. Based on the decrease in anogenital distance and delay in preputial separation, I conclude that the NOAEL and LOAEL for butylbenzyl phthalate from this study are 100 mg/kg-day and 500 mg/ kg-day, respectively. These data are not amenable to benchmark dose modeling because only one exposure level shows a response statistically different from control. Tyl et al. (2004) conducted a two-generation reproductive toxicity study in Sprague-Dawley rats (30 rats per sex per dose). Butylbenzyl phthalate was administered in the feed at 0, 750, 3750, or 11250 ppm (equivalent to 0, 50, 250, or 750 mg/kg-day as calculated by the study authors). The major effects noted in this study are summarized in Table 6. From these results the authors concluded that in rats under the conditions of this study:  The F0 and F1 parental systemic and reproductive NOAEL and the F1 and F2 offspring reproductive toxicity NOAELs are 3750 ppm (equivalent to approximately 250 mg/kg-day).

Table 5 Studies considered for the derivation of the reference dose for diisobutyl phthalate. Critical Effect

NOAEL mg/ kg-day

LOAEL mg/kgday

BMDL mg/ kg-day

Reference

Decreased fetal testosterone production

100

300

80

Decreased fetal testosterone; testicular histopathology; vacuolization of Sertoli cells; multinuclear gonocytes; reduced anogenital distance in male pups

Not established

600

Not calculated

Howdeshell et al. (2008) Borch et al. (2006a)

Howdeshell et al. (2008) is chosen over Borch et al. (2006a) because it has the lower NOAEL and LOAEL.

95

R. Benson / Regulatory Toxicology and Pharmacology 53 (2009) 90–101 Table 6 Summary of effects reported after exposure to BBP (Tyl et al., 2004). Exposure to 11,250 ppm (equivalent to 750 mg/kg-day)

Exposure to 11,250 ppm (equivalent to 750 mg/kg-day) Exposure to 3,750 ppm (equivalent to 250 mg/kg-day)

Exposure to 750 (equivalent to 50 mg/kgday)

F1 Adults Reduced mating and fertility indices; Reduced number of implantation and live pups per litter on PND 0; Reduced adult epididymal sperm motility and sperm concentration; Increased incidence of reproductive tract malformations from perinatal exposure; Increased lesions in adult male testes, epididymides, and prostate from perinatal exposure F1 and F2 Offspring Reduced body weights in pups during lactation; Shortened anogenital distance in males; Delayed acquisition of puberty in males and females; retention of nipples and areolas in preweaning males; Male reproductive tract malformations F1 and F2 Offspring Reduced anogenital distance in males at birth F1 Adults No effects on reproductive function, reproductive organ weights, or histopathology F1 adults, F1 and F2 Offspring No effects

 The F1 and F2 offspring reproductive toxicity NOEL (no observed effect level) in males is 750 ppm (equivalent to approximately 50 mg/kg-day) based on the significantly shortened anogenital distance in F1 and F2 male pups at birth at 3750 ppm and at 11,250 ppm, with no effects on reproductive development, structures, or functions at 3750 ppm. Tyl et al. (2004) concluded that the reduced anogenital distance in F1 and F2 male pups at 3750 ppm (equivalent to 250 mg/kg-day) was an effect level, but not an adverse effect level, because the anogenital distance was only decreased to 91.8% of control in F1 and 97.1% of control in F2. These authors did consider the decrease in anogenital distance to 83% of control in F1 and to 86.3% of control in F2 male pups at 11,250 ppm (equivalent to 750 mg/kg-day) to be an adverse effect level. I conducted a benchmark dose analysis of the data on anogenital distance presented in Table 3 of Tyl et al. (2004). The Hill model provided the best fit. The BMD for a 1 standard deviation decrease (BMD1SD) is 240 mg/kg-day in F1 male pups and 300 mg/kg-day in F2 male pups. The corresponding 95% lower confidence (BMDL1SD) is 130 mg/kg-day in F1 male pups and 210 mg/kg-day in F2 male pups. A more recent study (Howdeshell et al., 2008) measured testosterone production following administration of butylbenzyl phthalate. Pregnant Sprague-Dawley rats were treated from GD 8 to 18 by gavage with 0 (n = 9), 100 (n = 4), 300 (n = 5), 600 (n = 2), or 900 (n = 2) mg/kg-day. The day of a sperm plug positive is defined as GD 1. Ex vivo testosterone production in male fetuses from the litter was measured by radioimmunoassay at GD 18 using a 3 h incubation period as described by Wilson et al. (2004). Data were reported as litter mean and standard error. There was no effect on testosterone production at 100 mg/kg-day and a statistically significant decrease (p < 0.01) in testosterone production at 300 mg/kg-day and above. The NOAEL is 100 mg/kg-day and the LOAEL is 300 mg/kgday. I conducted a benchmark dose analysis of the data. The Linear model gave the best fit with a BMD1SD of 133 mg/kg-day and a BMDL1SD of 102 mg/kg-day. The studies considered for the derivation of the RfD for butylbenzyl phthalate are summarized in Table 7. Using the BMDL1SD

of 102 mg/kg-day for decreased testosterone production from Howdeshell et al. (2008) and a total uncertainty factor of 100 (10 each for interspecies and intraspecies extrapolation), the RfD for butylbenzyl phthalate is 1 mg/kg-day (rounded from 1.02 mg/kgday), equivalent to 0.00327 mmol/kg-day using a molecular weight of 312. 6. Diethylhexyl phthalate—hazard identification, dose– response, and RfD NTP-CERHR has provided a summary of the toxicological data on diethylhexyl phthalate (NTP-CERHR, 2005; Kavlock et al., 2006). The most pertinent study is a multigeneration continuous breeding study in rats (NTP, 2004). Diethylhexyl phthalate was administered in the diet at concentrations of 1.5 (control), 10, 30, 100, 300, 1000, 7500, or 10000 ppm to groups of 17 male and 17 female Sprague-Dawley rats. The 10000 ppm animals only completed the F1 generation and were terminated due to the inability to produce an F2 generation. Based on measured feed consumption, the F0 animals consumed 0.12, 0.78, 2.4, 7.9, 23, 77, 592, or 775 mg/kg-day; the F1 animals consumed 0.09, 0.48, 1.4, 4.9, 14, 48, 391, or 543 mg/kg-day; and the F2 animals consumed 0.1, 0.47, 1.4, 4.8, 14, 46, or 359 mg/kg-day. The critical effects identified in this study were small or absent reproductive organs in combined F1 and F2 non-breeding adult males. The data are summarized in Table 8. The expert panel identified 100 ppm (equivalent to 3–5 mg/kgday) as the NOAEL and 300 ppm (equivalent to 14–23 mg/kg-day) as the LOAEL. I conducted a benchmark dose analysis of these data using the incidence by litter of any reproductive organ abnormality and the average exposure of the F1 and F2 animals. The best fitting model was the log-logistic model and gave a BMD10 of 42 mg/kgday and a BMDL10 of 27 mg/kg-day. Andrade et al. (2006) conducted an extensive exposure-response study following in utero and lactational exposure to diethylhexyl phthalate. Female Wistar rats (11–16 at each exposure) were treated daily by gavage in peanut oil from GD 6 to lactational day 21. The day of mating is defined as GD 0. The exposures were 0.015, 0.045, 0.135, 0.405, 1.215, 5, 15, 45, 135, or 405 mg/kg-day.

Table 7 Studies considered for the derivation of the reference dose for butylbenzyl phthalate. Critical Effect

NOAEL mg/kg-day

LOAEL mg/kg-day

BMDL mg/kg-day

Reference

Decreased fetal testosterone production Reduced anogenital distance in male pups; delay in preputial separation Reduced anogenital distance in male pups

100 100 250b

300 500a 750c

102 Not calculated 130

Howdeshell et al. (2008) Nagao et al. (2000) Tyl et al. (2004)

Howdeshell et al. (2008) is chosen over Tyl et al. (2004) and Nagao et al. (2000) because it has the lowest NOAEL, LOAEL, and BMDL. a Anogenital distance decreased to 92.3% of control; preputial separation increased from 43.2 ± 1.5 days to 44.5 ± 2.3 days. b The decrease to 91.8% of control in F1 and to 97.1% of control in F2 was considered an effect level, but not an adverse effect level by Tyl et al. (2004). c Decreased to 83% of control in F1 and to 86.3% of control in F2.

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Table 8 Reproductive organ abnormalities in combined F1 and F2 non-breeding adult males after exposure to DEHP (NTP, 2004). DEHP, ppm in feed N

1.5 (39)

10 (36)

30 (39)

100 (41)

300 (45)

1000 (43)

7500 (30)

Abnormality in: Testis Epididymis Seminal vesicle Prostate Any reproductive organ

0 0 0 0 0

0 0 1 0 1 (1)

0 0 0 0 0

0 0 0 0 0

4 3 2 0 5 (4)

3 3 0 4 7 (5)

21 7 0 1 22 (14)

Data expressed as number of animals (litters) affected. Source: Table 23 from NTPCERHR (2005).

Nipple retention and reduced anogenital distance were only seen at an exposure of 405 mg/kg-day. Delayed preputial separation, examined on PND 33 and after, was observed at exposures of 15 mg/kg-day and above. There was no effect on testis descent when examined on PND 15 and after or on intratesticular testosterone at PND 1 at any exposure. Testis weight was statistically significantly increased at exposures of 5, 15, 45, and 135 mg/kg-day and decreased (not statistically significant) at 405 mg/kg-day. However, the increase was only greater than 10% at 135 mg/kg-day (I estimated the% change from the bar graphs in the publication). Accordingly, I do not consider the increase in testis weight to be biologically significant in this study. Histopathological examination of the testis on PND 1 showed the presence of bi and multinucleated gonocytes at exposures of 135 and 405 mg/kg-day. Histopathological examination of the testis on PND 22 showed signs of reduced germ cell differentiation in seminiferous tubules at exposures of 135 and 405 mg/kg-day. Based on the delay in preputial separation, the NOAEL in this study is 5 mg/kg-day and the LOAEL is 15 mg/kg-day. The data in the paper are not expressed in a way that is amenable to benchmark dose analysis. Several studies have investigated the effect of diethylhexyl phthalate on fetal testosterone. Andrade et al. (2006) did not find a decrease in testicular testosterone at PND 1. Borch et al. (2004, 2006b) reported statistically significant decreases in fetal testicular testosterone production ex vivo and in fetal testicular testosterone content at GD 21 at both 300 and 750 mg/kg-day. The day following mating is defined as GD 1. Parks et al. (2000) reported a decrease in testosterone production and reduced testicular testosterone from GD 17 to PND 2 at 750 mg/kg-day. The day of a sperm plug positive is defined as GD 1. A more recent study (Howdeshell et al., 2008) measured testosterone production following administration of diethylhexyl phthalate. Pregnant Sprague-Dawley rats were treated from GD 8 to 18 by gavage with 0 (n = 4), 100 (n = 4), 300 (n = 4), 600 (n = 4), or 900 (n = 4) mg/kg-day. The day of a sperm plug positive is defined as GD 1. Ex vivo testosterone production in male fetuses from the litter was measured by radioimmunoassay at GD 18 using a 3 h incubation period as described by Wilson et al. (2004). Data were reported as litter mean and standard error. There was no effect on testosterone production at 100 mg/kg-day and a statistically significant decrease (p < 0.01) in testosterone production at 300 mg/kg-day and

above. Accordingly, the NOAEL is 100 mg/kg-day and the LOAEL is 300 mg/kg-day. I conducted a benchmark dose analysis of their results. The Hill model provided the best fit with a BMD1SD of 142 mg/kg-day and a BMDL1SD of 67 mg/kg-day. Borch et al. (2006b) used quantitative RT-PCR and immunohistochemical staining to show that there was reduced mRNA expression and protein synthesis for SR-B1 (scavenger receptor class B-1), StAR, PRB (peripheral benzodiazepine receptor), and P450scc. These proteins are involved in key steps in testosterone synthesis in the fetal Leydig cell. The studies considered for the derivation of the RfD for diethylhexyl phthalate are summarized in Table 9. Based on the BMDL10 of 27 mg/kg-day for abnormalities in male reproductive organs (NTP, 2004) and an uncertainty factor of 100 (10 each for interspecies and intraspecies extrapolation), the RfD for diethylhexyl phthalate is 0.3 mg/kg-day (rounded from 0.27 mg/kgday), equivalent to 0.000692 mmol/kg-day using a molecular weight of 390. 7. Dipentyl phthalate—hazard identification, dose–response, and RfD Toxicological data on dipentyl phthalate is extremely limited (Foster et al., 1980; 1983; Howdeshell et al., 2008). Howdeshell et al., 2008 characterized the reproductive toxicity of dipentyl phthalate. Pregnant Sprague-Dawley rats were treated from GD 8 to 18 by gavage with 0 (n = 5), 25 (n = 5), 50 (n = 4), 100 (n = 6), 200 (n = 4), 300 (n = 2), 600 (n = 1), or 900 (n = 2) mg/kg-day. The day of a sperm plug positive is defined as GD 1. Maternal body weight gain was statistically significantly (p < 0.05) depressed at 300 mg/kg-day and above. There was also complete loss of litters at these exposures. Ex vivo testosterone production in male fetuses from the surviving litters was measured by radioimmunoassay at GD 18 using a 3 h incubation period as described by Wilson et al. (2004). Data were reported as litter mean and standard error. There was no effect on testosterone production at 25 or 50 mg/ kg-day and a statistically significant decrease (p < 0.01) in testosterone production at 100 and 200 mg/kg-day. The NOAEL is 50 mg/kg-day and the LOAEL is 100 mg/kg-day in this study. I conducted a benchmark dose analysis of the data. The Polynomial model gave the best fit with a BMD1SD of 26 mg/kg-day and a BMDL1SD of 17 mg/kg-day. Only the study of Howdeshell et al. (2008) is appropriate for the derivation of the RfD for dipentyl phthalate. Using the BMDL1SD of 17 mg/kg-day for decreased testosterone production (Howdeshell et al., 2008) and a total uncertainty factor of 100 (10 each for interspecies and intraspecies extrapolation), the RfD for dipentyl phthalate is 0.2 mg/kg-day (rounded from 0.17 mg/kg-day), equivalent to 0.000548 mmol/kg-day using a molecular weight of 310. 8. Diisononyl phthalate—hazard identification, dose–response, and RfD NTP-CERHR has provided a summary of the toxicological data on diisononyl phthalate (NTP-CERHR, 2003c; Kavlock et al.,

Table 9 Studies considered for the derivation of the reference dose for diethylhexyl phthalate. Critical Effect

NOAEL mg/kg-day

LOAEL mg/kg-day

BMDL mg/kg-day

Reference

Small or absent male reproductive organs in combined F1 and F2 Delay in preputial separation Decreased fetal testosterone production

3–5 5 100

14–23 15 300

27 Not calculated 67

NTP (2004) Andrade et al. (2006) Howdeshell et al. (2008)

NTP (2004) is chosen over Howdeshell et al. (2008) because it has a lower BMDL. NTP (2004) is chosen over Andrade et al. (2006) because it exposed animals for twogenerations and included more animals.

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2002c). The studies on reproductive and developmental toxicity showed that exposure to diisononyl phthalate during gestation can affect development of the kidneys and skeletal system of the fetus and can cause decreased birth weights. The studies reviewed by the expert panel reported no evidence of adverse effects on the developing reproductive system of rats. Since the review by the expert panel, two research groups (Gray et al., 2000; Borch et al., 2004) reported adverse effects on the developing male reproductive system following in utero exposure. Gray et al. (2000) treated pregnant Sprague-Dawley rats (6–10 animals) by gavage in corn oil with 0 or 750 mg/kg-day diisononyl phthalate (CAS # 68515-48-0) from GD14 to PND 3. The day of a sperm plug positive is defined as GD 1. Treatment did not cause maternal toxicity or reduced litter size. Diisononyl phthalate induced a statistically significant increase in retained areolas/nipples in 22% of male pups and a statistically significant increase in reproductive system malformations in 7.7% of male pups. The malformations in the reproductive system included small and atrophic testes; flaccid, fluid-filled testes; unilateral epididymal agenesis with hypospermatogenesis; and scrotal fluid-filled testis devoid of spermatids. There were no effects of diisononyl phthalate on testis weight, anogenital distance, age at preputial separation, hypospadias, or undescended testes. This study is limited because only one dose of diisononyl phthalate was tested and the incidence of adverse effects was low. Borch et al. (2004) investigated the effect of diethylhexyl phthalate, diisononyl phthalate, or a combination of these chemicals on hormone levels. Thirty-two pregnant Wistar rats were randomly assigned to four groups and exposed by gavage to peanut oil vehicle, 300 mg/kg-day diethylhexyl phthalate, 750 mg/kg-day diisononyl phthalate (CAS # 28553-12-0), or 300 mg/kg-day diethylhexyl phthalate plus 750 mg/kg-day diisononyl phthalate. Male fetuses were sacrificed on GD 21 and blood and testes were collected for hormone analysis [testosterone and plasma luteinizing hormone (LH)]. The day following mating is defined as GD 1. In this study there was a statistically significant reduction in testicular testosterone production and in testicular testosterone content in animals exposed to diethylhexyl phthalate, diisononyl phthalate, and diethylhexyl phthalate plus diisononyl phthalate. The testicular testosterone content was reduced to 30% of the control in the diethylhexyl phthalate group, to 26% of control in the diisononyl phthalate group, and to 16% of control in the diethylhexyl phthalate plus diisononyl phthalate group. Plasma testosterone was only significantly reduced in the diethylhexyl phthalate plus diisononyl phthalate group (to 48% of control). Plasma LH was statistically significantly increased in only the group treated with diethylhexyl phthalate plus diisononyl phthalate (190% of control). I estimated the% change from the bar graphs in the publication. This study is limited because only one group was exposed to diisononyl phthalate alone and only hormone levels were measured. The studies considered for the derivation of the RfD for diisononyl phthalate are summarized in Table 10. The data for diisononyl phthalate are limited because each study used only one exposure (750 mg/kg-day) and neither established a NOAEL. Unfortunately, Howdeshell et al. (2008) did not investigate the effect of diisononyl phthalate on fetal testosterone synthesis. Based on the retained areolas/nipples in male pups and malformations of the testes (Gray

et al., 2000) and the decreased testosterone levels in male fetuses (Borch et al., 2004), the LOAEL for diisononyl phthalate is 750 mg/kg-day. Using a total uncertainty factor of 1000 (10 for extrapolation from a LOAEL to a NOAEL, and 10 each for interspecies and intraspecies extrapolation), the RfD for diisononyl phthalate is 0.8 mg/kg-day (rounded from 0.75 mg/kg-day), equivalent to 0.00179 mmol/kg-day using a molecular weight of 419. 9. Summary of derived reference doses The RfDs are presented in Table 11. Each RfD includes an uncertainty factor of 10 each for interspecies and intraspecies extrapolation. The RfD for diisononyl phthalate also includes an additional uncertainty factor of 10 to extrapolate from a LOAEL to a NOAEL. Each of these phthalate esters causes a similar spectrum of adverse effects on the developing male reproductive system following perinatal exposure. The most likely underlying cause is the decrease in the concentration of the androgen receptor–testosterone complex due to a decrease in testosterone synthesis in fetal Leydig cells during the critical window in development (Reviewed in Foster, 2006; Hughes, 2001; Hughes et al., 2001). For dibutyl phthalate, diisobutyl phthalate, butylbenzyl phthalate, dipentyl phthalate, and diisononyl phthalate there is direct evidence for reduced testosterone in the fetal testis (Borch et al., 2004; 2006a; Hotchkiss et al., 2004; Howdeshell et al., 2007; 2008; Lehmann et al., 2004; Parks et al., 2000; Wilson et al., 2004). For dibutyl phthalate, diisobutyl phthalate, and diethylhexyl phthalate there are toxicogenomic data showing a decrease in the expression of mRNA and proteins for critical steps in steroidogenesis (Borch et al., 2006a; Lehmann et al., 2004; Howdeshell et al., 2007; Thompson et al., 2004; 2005). Liu et al. (2005) have shown that developmentally toxic phthalate esters (dibutyl phthalate, butylbenzyl phthalate, dipentyl phthalate, and diethylhexyl phthalate) were indistinguishable in their effects on global gene expression and that non-developmentally toxic phthalate esters (dimethyl phthalate, diethyl phthalate, dioctyl terephthalate) do not cause these changes in global gene expression. These chemicals cause the same spectrum of adverse effects on the developing male reproductive system by the same mechanism of action. Therefore, it is appropriate to derive relative potency factors (RPFs). Diethylhexyl phthalate is chosen as the index chemical and is assigned a relative potency of 1. The RPFs in Table 12 were calculated using the ratios of the unrounded RfDs based on mmol/ kg-day (Table 11). The exposure-response information used to derive the RPFs was taken from different laboratories using different exposure regimens. In addition, in some cases a different toxicological endpoint was used as the critical effect. Therefore, there is some inaccuracy in the RPFs presented in Table 12. 10. Cumulative risk assessment for simultaneous exposure Dibutyl phthalate, diisobutyl phthalate, butylbenzyl phthalate, diethylhexyl phthalate, dipentyl phthalate, and diisononyl phthalate cause a similar spectrum of adverse health effects in laboratory animals and operate by the same mechanism of action, reduction of testosterone in the fetal testes (Foster, 2006; Howdeshell et al., 2007; 2008). Therefore, it is appropriate to use a dose addition

Table 10 Studies considered for the derivation of the reference dose for diisononyl phthalate. Critical effect

NOAEL mg/kg-day

LOAEL mg/kg-day

BMDL mg/kg-day

Reference

Retained areolas/nipples in males; malformation of testes Decreased fetal testosterone

Not established Not established

750 750

Not calculated Not calculated

Gray et al. (2000) Borch et al. (2004)

Neither study established a NOAEL as only one exposure was studied. Neither study is clearly superior and are used as co-principal studies.

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Table 11 Reference dose for selected phthalate esters. Chemical

Critical effect and reference

POD/UF*

RfD

DBP

Decreased fetal testosterone with a NOAEL of 30 mg/kg-day Lehmann et al. (2004)

NOAEL/100

DiBP

Decreased fetal testosterone production with a BMDL1SD of 80 mg/kg-day Howdeshell et al. (2008)

BBP

Decreased fetal testosterone production with a BMDL1SD of 102 mg/kg-day Howdeshell et al. (2008) Small or absent male reproductive organs with a BMDL10 of 27 mg/kg-day NTP (2004)

BMDL1SD/ 100 BMDL1SD/ 100 BMDL10/ 100 BMDL1SD/ 100 LOAEL/ 1000

0.3 mg/kg-day or 0.00108 mmol/kg-day 0.8 mg/kg-day or 0.00288 mmol/kg-day 1 mg/kg-day or 0.00327 mmol/kg-day 0.3 mg/kg-day or 0.000692 mmol/kg-day 0.2 mg/kg-day or 0.000548 mmol/kg-day 0.8 mg/kg-day or 0.00179 mmol/kg-day

DEHP Dipentyl phthalate (DPP) DiNP

Decreased fetal testosterone production with a BMDL1SD of 17 mg/kg-day Howdeshell et al. (2008) Decreased fetal testosterone; Retained areolas/nipples with a LOAEL of 750 mg/kg-day Borch et al. (2004) and Gray et al. (2000)

POD is the Point of Departure and UF is the total uncertainty factor. RfD (mmol/kg-day) = RfD (mg/kg-day)/mol wt (mg/mmol)

Table 12 Relative potency factors.

Table 14 Hazard quotient and hazard index calculation for a US population.

Chemical

Relative potency factor#

DPP DEHP DBP DiNP DiBP BBP

1.26 1.00 0.64 0.39 0.24 0.21

# DEHP is assigned a relative potency factor of 1. The other values are calculated by dividing the RfD for DEHP in mmol/kg-day by RfD for the other chemical in mmol/kg-day from Table 9. The values are rounded to two significant digits.

model (Lambert and Lipscomb, 2007; Kortencamp, 2007) in a cumulative risk assessment for these phthalate esters. A response addition model is not appropriate. To do a cumulative risk assessment, it is necessary to have an RfD (see previous section) and some information on exposure. For the purposes of illustration, exposure information of Kohn et al. (2000) for a US population (see Table 13) and of Wittassek and Angerer (2008) for a German population (see Table 15) will be used. There is no exposure information available for dipentyl phthalate for either population. Although the exposure reconstruction presented by Kohn et al. (2000) and by Wittassek and Angerer (2008) is intended to capture the typical range of exposure of the population examined, these data may not capture extremes of exposure. For example, Hauser

Table 13 Exposure to selected phthalate esters for a US population. Chemical

Median exposure (mg/kg-day)

95th Percentile exposure (mg/kg-day)

Maximum exposure (mg/kg-day)

DBP DiBP BBP DEHP DPP DiNP

0.0013 0.0002 0.00088 0.00071 Not Available
0.0061 0.0011 0.0040 0.0036 Not Available 0.0017

0.094 (0.26)* 0.016 0.029 0.046 Not Available 0.022

Exposure to BBP, DEHP, and DiNP is directly from Kohn et al. (2000), Table 2 (n = 289). The original data used by Kohn et al. (2000) reported the urinary excretion of DBP and DiBP as a single summed value. Subsequent data collected by NHANES (DHHS, CDC, 2005) shows that urinary excretion of DiBP (creatinine corrected) is approximately 15% of the excretion of DBP (creatinine corrected). Accordingly, the DBP values reported by Kohn et al. (2000) have been reduced by 15% and the remainder is attributed to DiBP. No publication has updated the exposure reconstruction using the more recent NHANES data (DHHS, CDC, 2005). If these more recent data were used for an exposure reconstruction using the method of Kohn et al. (2000), comparable exposures to those presented in the table would be obtained. * Based on Hauser et al. (2004). See text in Section 10.

Chemical DBP DiBP BBP DEHP DPP DiNP

Median exposure

95th percentile exposure

Maximum exposure

hazard quotient 0.004 0.0003 0.0009 0.002 Not calculated Not calculated Hazard index 0.007

hazard quotient 0.02 0.001 0.004 0.01 Not calculated 0.002 Hazard index 0.04

hazard quotient 0.3 (0.9)* 0.02 0.03 0.2 Not calculated 0.03 Hazard index 0.6 (1)**

The hazard quotient is the exposure divided by the RfD. The hazard index is the sum of the individual hazard quotients. All values rounded to 1 significant digit. The hazard quotients for dipentyl phthalate are not calculated because no human exposure information is available. The hazard quotient for the median exposure to diisononyl phthalate is not calculated because the median urinary concentration was below the limit of detection. * Using calculated exposure data from Hauser et al. (2004). See text in Section 10. ** Using the hazard quotient calculated from Hauser et al. (2004) for dibutyl phthalate and the same values for maximum exposure to the other phthalate esters.

Table 15 Exposure to selected phthalate esters for a German population. Chemical

Median exposure (mg/kg-day)

95th Percentile exposure (mg/kg-day)

Maximum exposure (mg/kg-day)

DBP DiBP BBP DEHP DPP DiNP

0.0021 0.0015 0.0003 0.0027 Not available 0.0006

Not Not Not Not Not Not

0.230 0.0273 0.0022 0.0422 Not available 0.0368

provided provided provided provided available provided

Data are from Wittassek and Angerer (2008) for 102 individuals. The authors calculated the oral exposure from the urinary levels of the metabolites and urinary excretion factors determined from human metabolism studies.

et al. (2004) report on a single individual who was exposed to dibutyl phthalate from a pharmaceutical containing dibutyl phthalate in the delayed release coating of the product. The individual’s urinary monobutyl phthalate concentration was 16,868 ng/mL (6,180 lg/g creatinine), approximately two orders of magnitude greater than the 95th percentile of the US population reported in the 1999–2000 NHANES data set. Although not reported by Hauser et al. (2004), I calculated this individual’s oral exposure to dibutyl phthalate to be 0.26 mg/kg-day. This calculation used the equation in David (2000), information on the excretion of the monobutyl phthalate relative to dibutyl phthalate ingested (0.69) from Anderson et al. (2001), an excretion rate of creatinine of 23 mg/kg-day (Kohn et al., 2000), and the urinary monobutyl phthalate concentration reported by Hauser et al. (2004).

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The two common approaches for a cumulative risk assessment using dose addition are the hazard index and the relative potency approaches (Lambert and Lipscomb, 2007). The step-by-step calculation of the hazard index approach for the US population is provided below. 1. Calculate the hazard quotient for the median exposure to each phthalate ester by dividing the median exposure in mg/kg-day to each phthalate ester from Table 13 by the reference dose in mg/kg-day for the respective phthalate ester from Table 11. The resulting hazard quotient is in the Median Exposure column in Table 14. 2. Add the hazard quotients in the Median Exposure column in Table 14 to get the hazard index for the median exposure. 3. Repeat steps 1 and 2 for the 95th Percentile Exposure and Maximum Exposure. The results are presented in Table 14 The same procedure is used to calculate the hazard index for the German population from the exposure information in Table 15. The results for the German population are presented in Table 16. The step-by-step calculation of the hazard index using the relative potency approach for the US population is provided below. 1. Calculate the exposure in DEHP equivalents (mmol/kg-day) for the median exposure to each phthalate ester by dividing the median exposure in mg/kg-day to each phthalate ester from Table 13 by the molecular weight of the respective phthalate ester (Table 1). 2. Multiply the result from step 1 by the relative potency factor from Table 12. 3. Add the values calculated in step 2. This is the total exposure in DEHP equivalents. 4. Divide the total exposure in DEHP equivalents from step 3 by the RfD for DEHP from Table 11 (0.000692 mmol/kg-day). The resulting value is the hazard index for the median exposure. 5. Repeat steps 1–4 for the 95th Percentile Exposure and Maximum Exposure. The hazard indexes calculated using the relative potency approach for the US and German populations are identical to the hazard indexes presented in Tables 14 and 16. Therefore, separate tables with the results using the relative potency approach for the US and German populations are not presented. A HI of less than 1 means that there is little probability that an adverse effect might be observed from the exposure to the chemicals. A HI of 1–100 means that there is a potential that adverse effects might be observed. A HI of 100 or more indicates that the measured exposure is comparable to the NOAEL and LOAEL observed in laboratory animals and indicates a high potential that

Table 16 Hazard quotient and hazard index calculation for a German population. Chemical

DBP DiBP BBP DEHP DPP DiNP

Median exposure

Maximum exposure

Hazard quotient

Hazard quotient

0.007 0.002 0.0003 0.009 Not calculated 0.0008 Hazard index 0.02

0.8 0.003 0.002 0.01 Not calculated 0.05 Hazard index 0.8

The hazard quotient is the exposure divided by the RfD. The hazard index is the sum of the individual hazard quotients. All values rounded to 1 significant digit. The hazard quotients for dipentyl phthalate are not calculated because no human exposure information is available.

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adverse effects might be observed in humans at these exposures. As can be seen from the results in Tables 14 and 16, the HI for the median, 95 percentile, and maximum exposed individual are equal to or less than 1. Thus, it is unlikely that humans are suffering adverse developmental effects from current environmental exposure to these phthalate esters. 11. Summary This paper is the first to derive RfDs for the phthalate esters based on adverse developmental effects in rats following in utero exposure. The RfDs are summarized in Section 9 and Table 11. For dibutyl phthalate, diisobutyl phthalate, butylbenzyl phthalate, dipentyl phthalate, and diisononyl phthalate, the critical effect is decreased fetal testosterone near the end of the gestation period. For diethylhexyl phthalate the critical effect is small or absent male reproductive organs following in utero exposure. Although not used as the critical effect, diethylhexyl phthalate also decreases fetal testosterone near the end of the gestation period at a somewhat higher exposure (BMDL10 of 27 mg/kg-day for malformations in NTP, 2000, and BMDL1SD of 67 mg/kg-day for decreased testosterone synthesis in Howdeshell et al., 2008). It is possible that if Howdeshell et al. (2008) had studied the effect of diethylhexyl phthalate on testosterone synthesis later in the gestation period, decreased fetal testosterone might have been the critical effect for diethylhexyl phthalate. The assessment for diisononyl phthalate would be improved by having a complete exposure-response study for decreased fetal testosterone below the exposure of 750 mg/kg-day. This chemical was not included in the study by Howdeshell et al. (2008). The RfDs for these phthalate esters would also benefit from the development of physiologically based pharmacokinetic models (Clewell et al., 2008) for the rat and human to better assess the interspecies and intraspecies uncertainty factors and allow derivation of the RfD from internal dose, rather than external exposure. The assessments in this paper use default values of 10 each for the interspecies and intraspecies uncertainty factors because there are no robust data to use to depart from the default values. This paper is also the first to provide methods for conducting a cumulative risk assessment for simultaneous exposure to dibutyl phthalate, diisobutyl phthalate, butylbenzyl phthalate, diethylhexyl phthalate, dipentyl phthalate, and diisononyl phthalate. The methods presented include the hazard index approach based on the RfD and exposure information in people on the phthalate ester of interest and the relative potency approach based on calculation of diethylhexyl phthalate equivalents from exposure information in people for each phthalate ester of interest. Sample cumulative risk assessments are provided for a US and a German population using exposure information published by other researchers (Kohn et al., 2000; Wittassek and Angerer, 2008). These results are presented in Tables 14 and 16 and show that it is unlikely that humans are suffering adverse developmental effects from current environmental exposure to these phthalate esters. These approaches to a cumulative risk assessment can be used when new exposure information is available or when new studies show that additional phthalate esters cause developmental toxicity due to decreased testosterone synthesis in the fetal testes. As the results for the hazard index and relative potency approaches give the same results, I suggest using the less complex hazard index approach for a future risk assessment. Although the focus of this paper is on the reduction in fetal testosterone and subsequent malformations of the male reproductive tract caused by exposure to phthalate esters, simultaneous exposure to androgen receptor antagonists such as vinclozolin, flutamide, procymidone, linuron, or other inhibitors of androgen

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synthesis such as prochloraz will enhance the male reproductive tract malformations induced by the phthalate esters (Christiansen et al., 2008; Rider et al., 2008). Christiansen et al. (2008) showed that a mixture of vinclozolin, flutamide, and procymidone act together in an additive way to induce male reproductive tract malformations. Rider et al. (2008) have shown that a mixture of vinclozolin, procymidone, linuron, prochloraz, butylbenzyl phthalate, dibutyl phthalate, and diethylhexyl phthalate act together to disrupt male reproductive tract differentiation and induce malformations. The effects observed for this mixture of chemicals followed a dose addition model. It is beyond the scope of this paper to provide detailed exposure-response information for these other chemicals. 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