Regulatory Toxicology and Pharmacology 54 (2009) 134–142

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Regulatory Toxicology and Pharmacology

journal homepage: www.elsevier .com/locate /yr tph

The role of developmental toxicity studies in acute exposure assessments:Analysis of single-day vs. multiple-day exposure regimens

Allen Davis a,b,*, Jeff S. Gift b, George M. Woodall b, Michael G. Narotsky c, Gary L. Foureman b

a Oak Ridge Institute for Science and Education, Research Triangle Park, NC 27711, USAb United States Environmental Protection Agency, National Center for Environmental Assessment, 109 TW Alexander Dr, Mail Code B243-01, Research Triangle Park, NC 27711, USAc United States Environmental Protection Agency, National Health and Environmental Effects Research Laboratory, Research Triangle Park, NC 27711, USA

a r t i c l e i n f o a b s t r a c t

Article history:Received 25 September 2008Available online 21 March 2009

Keywords:Developmental toxicityButylbenzyl phthalateTributyltinBenchmark doseRisk assessmentAcute exposureSingle-dayMultiple-day

0273-2300/$ - see front matter Published by Elsevierdoi:10.1016/j.yrtph.2009.03.006

* Corresponding author. Address: United StateAgency, National Center for Environmental AssessmMail Code B243-01, Research Triangle Park, NC 27711

E-mail address: davis.allen@epa.gov (A. Davis).

In accordance with most toxicity guidelines, developmental studies typically utilize repeated exposures,usually throughout gestation or during organogenesis in particular. However, it is known that develop-mental toxicity may occur in response to single exposures, especially during specific windows of suscep-tibility. An overview of the available literature gave sufficient evidence that for many agents, the sameendpoints observed in repeated dose, multiple-day studies were also observed in single-day exposures,thus indicating the relevance of developmental toxicity to health assessments of acute exposures. Fur-ther, results of benchmark dose modeling of developmental endpoints indicated that for embryo lethal-ity, single-day exposures required a two- to fourfold higher dose than the multiple-day exposures toproduce the same level of response. For fused sternebrae, exposures on specific days produced equivalentlevels of response at doses that were more similar to those utilized in the repeated exposures. Apprecia-ble differences in biological half-life (and corresponding dose metrics) as well as specific windows of sus-ceptibility may partially explain the observed multiple- vs. single-day exposure dose–responserelationships. Our results highlight the need of a more thorough evaluation of outcomes from repeateddose developmental toxicity studies in regards to their importance to chronic and acute risk assessments.

Published by Elsevier Inc.

1. Introduction

Prenatal developmental toxicity studies are designed to providegeneral information concerning the effects of exposure to the preg-nant test animal on the developing organism (US EPA, 1998).Although exposures during a typical guideline prenatal develop-mental toxicity study are designed to only include either the entireperiod of gestation or limited species-specific gestation periods(typically only 10–20 days for murine species), developmentalendpoints are considered to be an integral concern in the assess-ment of potential health effects from continuous lifetime expo-sures to a toxicant. Pregnancy and fetal development are thusconsidered to represent a potentially susceptible life stage thatshould be considered in lifetime or chronic assessments. If studiesand information on reproductive or developmental toxicity are ab-sent in a health effect assessment, specific uncertainty factors (e.g.,database) may be applied to the final point of departure used inrisk calculations (US EPA, 1991; Chou et al., 1998).

Inc.

s Environmental Protectionent, 109 TW Alexander Dr,, USA.

It is also well established, however, that developmental toxicitymay occur in response to single exposures, such as during specificdevelopmental windows of susceptibility. Whereas this circum-stance does not influence the relevance of typical guideline devel-opmental studies to the evaluation of chronic or lifetimeassessments of health effects, it does signify that developmentalendpoints observed in these repeated dose studies are relevant inhealth effect assessments of shorter term exposures, includingacute exposures. Acute studies are generally defined as a singleexposure of 24 h or less in duration.

This work is an initial investigation into the issues entailed inutilizing standard developmental toxicity studies in the assess-ment of responses from an acute exposure scenario. Aspects con-sidered include a general query into the extent of the informationavailable documenting developmental effects from single-dayexposures and whether or not specific endpoints are observedin both the single-day exposure and the corresponding multi-ple-day exposure scenarios. As typically a multiple-day guidelinestudy only would be available for an acute exposure assessment,some quantitative information on the relationship between effec-tive doses for these scenarios is also required. Therefore, this workalso proposes and demonstrates an approach to address this crit-ical issue of dose relationships between typical multiple-day andsingle-day exposure scenarios in eliciting fetal toxicity. This ap-

A. Davis et al. / Regulatory Toxicology and Pharmacology 54 (2009) 134–142 135

proach was extended to provide insight into possible co-variables,particularly pharmacokinetics by incorporating and comparingdose relationships between agents that had contrasting half-lives in anticipation of revealing relationships of dose measuresthat may influence the single-day to multiple-day doserelationship.

1 Extra risk is calculated as: [P(d) � P(0)/[1 � P(0)] where P(d) is the response adose d and P(0) is the background risk at zero dose.

2 Relative deviation is calculated as: P(0) ± (P(0) � BMR) where P(0) is the modeestimated background response and the direction of the dose–response determineswhether the product of the background response and the BMR is added to orsubtracted from the background response.

2. Methods

2.1. Literature search

Searches were performed on PubMed and the TOXNET site(including DART and TOXLINE) using a range of keywords andcombinations, and were not limited by date of publication, species,chemical, or study design. For the purposes of this work, ‘‘acute”(i.e., single-day) exposures were defined as single exposures lastingup to 24 h.

For those substances for which acute developmental studies ofsufficient quality were identified, information on the properties ofthe chemical agents (e.g., half-life and major metabolites) were ob-tained from literature searches, monographs (e.g., ATSDR profiles),and additional databases (Hazardous Substances Data Bank(HSDB)). For the identified substances that had suitable, high-qual-ity single-day exposure developmental data available, paired mul-tiple-day exposure developmental studies were located utilizingthe same search criteria and strategy. In an effort to limit sourcesof variability, multiple-day exposure studies were matched ascompletely as possible to the paired single-day studies for speciesand strain, chemical form, exposure route and endpoint evaluationprotocol. Chemicals with contrasting half-lives (<24 or >24 h) werealso sought out and selected in order to examine the dose–re-sponse profiles of toxicants expected to accumulate readily in thebody, vs. those expected to clear rapidly.

2.2. Dose–response modeling of paired (single-day and multiple-day)studies

The Environmental Protection Agency’s (EPA) Benchmark DoseSoftware (BMDS, available at http://cfpub.epa.gov/ncea/cfm/recor-display.cfm?deid=164443) version 2.0 was used to model dose–response relationships. BMDS comprises multiple dichotomous,continuous, and nested models, and is normally used in riskassessment for the computation of benchmark doses (BMDs)and their lower confidence limits (BMDLs), based on a givenbenchmark response (BMR). For the purpose of this work, onlyBMDs (i.e., the maximum likelihood estimates or MLEs), and notthe BMDLs, were used as a measure of where a given response le-vel occurred on a dose–response curve. This strategy allowed forcomparative analysis at the same (or equieffective) response level,based on the MLEs for both single- and multiple-day dosingregimens.

Although there is no absolute basis to assume that the shape ofthe dose–response curve is similar across the different dosing reg-imens, a single model was chosen for each individual endpoint forthe sake of simplicity. Model choice was based on the Akaike Infor-mation Criterion (AIC) (Akaike, 1973; Linhart and Zucchini, 1986;Stone, 1998). The model that had the lowest summed AIC for theregimens containing the most dose groups and that met the fit cri-teria (p > 0.1 and/or visual inspection) was considered the mostsuitable model for all data sets. This model was then used to fitthe observed data for the remaining data sets that had fewer dosegroups.

The entire suite of dichotomous and continuous models avail-able in BMDS v2.0 was considered for this analysis. For dichoto-mous endpoints, the benchmark response (BMR) level used in

this analysis is 0.25 extra risk (ER).1 ER accounts for the spontane-ous background risk, and gives a standardized response across dos-ing regimens. Because the ED50 can usually be estimated moreprecisely, especially for dichotomous data (Finney, 1971) a BMRof 0.5 was originally the goal for this work. Data limitations, mainlydose–responses with insufficient response rates, made it necessaryto decrease the BMR down to 0.25.

For continuous endpoints (or dichotomous endpoints modeledas a continuous variable), the BMRs used in these analyses werebased on varying levels of relative deviation (RD)2 from the modelestimate of the control mean (i.e., background). When modeling eachcontinuous endpoint, the control groups for all dosing regimenswere combined under the assumption that the experimental vari-ability for the control groups was equal across studies. For all mod-els, it was assumed that the variance was non-homogenous (i.e., notconstant with increasing dose) and the models were allowed to esti-mate the variance for each dosing regimen.

In certain instances (e.g., embryo lethality), fetal incidence datawere used for analyses without regard to litter-specific data. This isunderstood to be an over-simplification as the litter is recom-mended as the unit of analysis in developmental toxicology studies(Haseman and Hogan, 1975). Dose–response software such asBMDS do offer nested models that take important statistical con-siderations such as litter specific covariates and intralitter correla-tions into account (i.e., the tendency for littermates to respondsimilarly to one another as compared to other litters in a dosegroup). This study did not take advantage of such nested modelsas (1) individual animal data (required for existing versions ofnested models) were not easily available and (2) the available fetalincidence data were determined to be sufficient to provide unbi-ased estimates of the average incidence of effects at different doses(Auton, 1994). It is acknowledged that by not considering intralit-ter correlation in the modeling of developmental toxicity data, thevariance estimates (and thus the confidence limits on calculatedBMDs) would be too small. Our analysis avoids this concern byusing the central estimate of response, the BMDs, rather than theBMDLs, in its comparison across dosing regimens. Additionally,BMDs and BMDLs were compared only with embryo lethality in or-der to evaluate the potential for underestimating the variance bycalculating and then comparing Rao–Scott transformations of thedata to the untransformed data (see below).

The relationships of the doses estimated to result in the sameresponse (i.e., the same BMR) were viewed as ‘‘relative potency”in this analysis, such that the lower the BMD calculated, the higherthe potency and vice-versa. Once BMDs at the same designatedBMR were generated for both dichotomous and continuous end-points, the results for the single-day exposures were compared tothe multiple-day exposures by calculating the ratio of the equief-fective BMDs (i.e., single-day BMD/multiple-day BMD). In orderto visually compare the shape of the dose–response curves for alldosing regimens for a particular endpoint, the model-estimateddose–response curves and observed data for individual dosing reg-imens were plotted together on one graph.

2.3. Dose–response of endpoints

Embryo lethality, reported as resorbed and dead fetuses, post-implantation loss, intra-uterine death, or prenatal death rate (i.e.,number of prenatal deaths/total number of implants) was modeled

t

l

136 A. Davis et al. / Regulatory Toxicology and Pharmacology 54 (2009) 134–142

as both a continuous and dichotomous endpoint for the selecteddata sets. When modeled as a continuous endpoint, the meansand standard deviations as reported in the studies were used,and the BMRs used were 2.0, 3.0, and 4.0 RD. In order for embryolethality to be modeled as a dichotomous endpoint, the reportedlitter means and standard deviations for both implantations andresorbed/dead fetuses were transformed into incidence data.

To illustrate the degree to which intralitter correlation mayhave influenced either BMD or BMDL estimates, Rao–Scott trans-formations were calculated and applied to dichotomized fetal inci-dence data. The Rao–Scott transformation allows one to treatcorrelated binomial random variables (such as effects within thesame litter) as if they were uncorrelated by scaling the numberof responses and total number on test by the design effect3: the ra-tio of the true variance to the variance under correlation (Rao andScott, 1992; Krewski and Zhu, 1995). Various levels and patterns ofassumed intralitter correlation (U) were examined including 0.1,0.4, and 0.7 (i.e., low, medium, and high) constant correlation acrossdose groups as well as two ‘‘graded” correlation scenarios where Uincreased with dose level. These transformed incidence data werethen modeled using the same model and BMRs as the untransformedincidence data. Transformed BMDs, and their corresponding lowerconfidence limits (the BMDLs), were then compared to the untrans-formed BMD or BMDL values to determine how controlling for var-ious assumed levels and patterns of U altered the modeling resultsacross dosing regimens.

Fetal weight reduction was modeled for both males and femalesas a continuous endpoint. The studies included in the analysis sta-ted that the litter was the unit of statistical analysis and thus thesummary data for fetal weight provided in the papers are litterestimates. Using these litter estimates and a BMR not based onstandard deviation avoided the above mentioned issues of intralit-ter correlation. The BMRs used for modeling fetal weight were 0.05,0.10, and 0.15 RD. Cleft palate and fused sternebrae were modeledas dichotomous endpoints using litter data as presented in thestudies included in this analysis. Available information allowedboth endpoints to be modeled as litter incidence thereby avoidingissues surrounding intralitter correlation.

3. Results

3.1. Identification of single- and multiple-day developmental studies

The original literature search of PubMed, TOXLINE, and DARTusing keywords to capture toxic outcomes (e.g., fetotoxicity, mal-formation, teratogenicity) resulted in the identification of 133 sin-gle-day exposure developmental toxicity studies investigating 58chemicals. Multiple classes of chemicals were represented by these133 studies: solvents, metals, plasticizers, alkylating agents, pesti-cides, and fungal and plant toxins.

The range of adverse developmental effects reported by theidentified single-day studies was comprehensive and not qualita-tively different from those reported in the multiple-day studies.Adverse outcomes included: embryo lethality (dead and resorbedfetuses, post-implantation loss, pre-implantation loss), reducedfetal weight, cleft palate, exencephaly, hydrocephaly, urogenitalmalformations, altered postnatal neurobehavioral development,and abnormal pubertal development and sexual behavior. Altera-tions in maternal effects such as weight gain, food consumption,clinical observations, and mortality were also reported by manystudies.

3 The design effect is: D = 1 + (m � 1)U, where m is the average litter size(calculated using the control litter sizes for the studies used in this analysis and Uis the assumed intralitter correlation (Williams, 1982).

3.2. Identification of paired studies

After appropriate and corresponding single- and multiple-daystudies were identified, two chemicals were chosen for further,more definitive analysis on the basis of quality of single- and mul-tiple-day exposure-response data, half-life, and environmental rel-evance. Butyl benzyl phthalate (BBP) was chosen to berepresentative of chemicals with a short (624 h) half-life relativeto the exposure regimens under consideration; the terminal half-life of mono-phthalate metabolites is 5.5–6.8 h, with virtually allmetabolites cleared within 24 h (Eigenberg et al., 1986). Tributyltinchloride (TBT) was chosen to be representative of chemicals with acorrespondingly long (P24 h) half-life that could be anticipated toexhibit dose accumulation to a much greater degree than the shorthalf-life chemical; estimates for half-lives range from 23 to 30 days(tissue not specified) for tributyltin (WHO IPCS, 1994).

The studies used in the analyses of both BBP and TBT used theWistar rat as the experimental animal model and employed oralgavage as the route of exposure (Ema et al., 1992, 1993, 1995,1997, 1999; Itami et al., 1990). The multiple-day studies exposedanimals from gestational day (GD) 7 to 15, corresponding to theperiod of major organogenesis in the rat. Results from single-dayexposures occurring on GD 13, 14, or 15 were available for BBPand on GD 7, 8, or 9 for TBT. In addition to single- and multiple-day exposures, studies were identified that exposed dams on 3consecutive days. These exposures, herein termed ‘‘intermediate-day” occurred on GD 7–9, 10–12, or 13–15 for both BBP and TBTand were originally designed by these authors to identify and nar-row the specific period most susceptible to the developmental ef-fects resultant from exposure to these particular agents. In all casesfor the studies utilized, dams were sacrificed on GD 20 and rele-vant developmental toxicological assessments performed.

The dose–response data was abstracted from these studies andused in the comparative analyses. Analysis of results from most ofthese studies indicated relatively minor incidences of maternaltoxicity and, as such, does not pose any challenge to interpretationof these data (as per US EPA developmental guidelines, US EPA,1991). In a few instances, however, considerable maternal toxicitywas noted in the form of mortality and decreased weight gain.

3.3. Modeling and analysis

3.3.1. Embryo lethalityThe log-Probit model was the model choice for BBP, with Fig. 1

showing the modeled dose–response curves for multiple-, inter-mediate-, and single-day exposures. Exposure on multiple-days(GD 7–15) resulted in the steepest dose–response curve, and wasthe only dosing regimen that resulted in complete (100%) embryolethality at the highest dose. The general shapes of the intermedi-ate-day (GD 7–9, 10–12, or 13–15) dose–response curves weresimilar to each other and to that of multiple-day exposure. Theseintermediate dosing regimens were judged less potent than themultiple-day exposure, exhibiting less than complete embryolethality (62–88%) at the highest dose levels. Exposures on sin-gle-days (GD 13, 14, or 15) displayed similar but more shallow re-sponse profiles than the longer exposures with potency rangingfrom 48% embryo lethality for exposure on GD 13 to almost 75%on GD 15.

The modeled BMD values for embryo lethality as a dichotomousendpoint at a uniform response rate (BMR of 0.25 ER) for all BBPresponse curves in Fig. 1 are shown in Table 1. The BMD calculatedfor BBP exposure on GD 7–15 was 586.4 mg/kg, the lowest andtherefore most potent of the exposure regimens examined for thisendpoint. The other BMDs for BBP were solved and expressed as aratio to the GD 7–15 BMD. The dose ratios associated with thesame response level (0.25) for the various single GD exposures

Fig. 1. Results from dose–response modeling of embryo lethality as a dichotomous variable for both butylbenzyl phthalate (BBP) and tributyltin (TBT). Observed data aredenoted by symbols as labeled in the figure; closed symbols represent statistically significant dose–responses (p < 0.05 or 0.01) and open symbols represent non-significantdose–responses as reported by Ema et al. (1992, 1993, 1995, 1997, 1999) and Itami et al. (1990); lines represent model results. BBP was fit by the log-Probit model; themodeled dose–response curves fit the experimental data acceptably for the multiple- and single-day exposures using both calculated p-values (for GD 7–15, p = 0.53) andvisual inspection (for single-day exposures) as criteria for goodness-of-fit, but appeared comparatively less accurate when fitting the intermediate-day exposures, with p-values < 0.10 for GDs 7–9 and 13–15. TBT was fit by the Weibull model; the dose–response curves fit all the experimental data acceptably for all exposures, as judged byvisual inspection and acceptable p-values as criteria for goodness-of-fit.

Table 1BMD calculations for embryo lethality for butyl benzyl phthalate and tributyltin,using dichotomized fetal incidence data.

Butyl Benzyl Phthalate Tributyltin

GD BMDa BMD Ratio to GD 7–15 GD BMDb BMD Ratio to GD 7–15

13 1231.8 2.101 7 53.9 3.56914 1040.4 1.774 8 60.8 4.02615 1140.8 1.945 9 68.5 4.5367–9 687.0 1.172 7–9 18.1 1.19910–12 756.5 1.290 10–12 83.6 5.53613–15 692.0 1.180 13–15 >100 >6.6237–15 586.4 1.000 7–15 15.1 1.000

a Log-Probit model, restricted to power P 1, BMR = 0.25 extra risk.b Weibull model, restricted to power P 1, BMR = 0.25 extra risk.

A. Davis et al. / Regulatory Toxicology and Pharmacology 54 (2009) 134–142 137

and the multiple GD 7–15 exposure were found to be nearly dou-bled; 2.1 for GD 13, 1.8 for GD 14 and 1.9 for GD 15. The ratios con-structed for the intermediate-day exposures were found to all bebetween the multiple- and single-day exposures: 1.2 for GD 7–9,1.3 for GD 10–12, and 1.2 for GD 13–15.

The Weibull model was the model choice for the TBT embryolethality data, also shown in Fig. 1. As with BBP, exposure to TBTon GD 7–15 resulted in the steepest dose–response curve andwas the only dosing regimen to result in complete (100%) embryolethality at the highest dose. Unlike BBP, however, intermediate-day exposures to TBT were very different from one another. Expo-sure to TBT on GD 7–9 resulted in a dose–response curve very sim-ilar to GD 7–15 in shape and potency, achieving 90% embryolethality. Exposure on GD 10–12 showed minimal response except

138 A. Davis et al. / Regulatory Toxicology and Pharmacology 54 (2009) 134–142

at the highest dose (approximately 45% incidence) and exposure onGD 13–15 did not increase embryo lethality over background lev-els (roughly 10%). Single-day exposures to TBT on GD 7, 8, or 9 re-sulted in similar dose–response curves that were approximatelylinear. The incidence of embryo lethality for single-day exposuresranged from 69% on GD 7 to 93% on GD 9.

The modeled BMD values for embryo lethality as a dichotomousendpoint at a uniform response rate (BMR of 0.25 ER) for all TBTresponse curves in Fig. 1 are shown in Table 1. The BMD calculatedfor TBT exposure on GD 7–15 was 15.1 mg/kg day, and like BBP, thelowest and therefore most potent of the exposure regimens exam-ined for this endpoint. As with BBP, the BMDs calculated for theother dosing regimens were calculated and expressed as a ratioto the GD 7–15 value. The BMD ratios for the various single GDexposures and the multiple GD 7–15 exposure were more thandoubled; 3.6 for GD 7, 4.0 for GD 8 and 4.5 for GD 9. The ratios con-structed for the intermediate-day exposure regimens were in con-trast with one another. Two regimens, GD 10–12 and 13–15,yielded ratios of 5.5 and >6, respectively, both as great as the sin-gle-day regimens whereas the remaining GD 7–9, yielded a ratio of1.2, quite similar to the GD 7–15 multiple-day exposure.

When the dichotomized embryo lethality data sets for BBP wereadjusted using the Rao–Scott transformation and modeled, the ra-tios of the transformed vs. untransformed BMDs did not changeappreciably across the different levels and patterns of assumedintralitter correlation (see Table 2). The differences between thetransformed and untransformed BMDs did not increase or decreasewith any particular pattern.

The ratios of transformed vs. untransformed BMDLs, however,varied greatly depending on the level and pattern of assumedintralitter correlations. Whereas the ratios of transformed anduntransformed BMDs never exceeded 3%, differences betweentransformed and untransformed BMDLs approached 50% in somecases. The greater the assumed intralitter correlation was (for bothpatterns, constant and increasing with dose), the lower the trans-formed BMDL was compared to the untransformed BMDL. Duethe very small differences between the various adjusted and unad-justed BMD ratios, indicating that correcting for assumed intralit-ter correlation did not change the estimated BMDs greatly, andthe large differences between BMDLs, it was concluded that forthe particular measure used in the main dose–response analyses(the BMD), using a model that did not take intralitter correlationinto account would make little difference to the outcomes of theanalysis.

When the dichotomized embryo lethality data sets for TBT wereadjusted using the Rao–Scott transformation and modeled as perBBP, the differences between transformed and untransformedBMDs were slightly larger than the differences seen with BBP (data

Table 2BMD and BMDL calculations and ratios for dead and resorbed fetuses for butyl benzyl ph

GD Assumed intralitter correlation

Low (Ua = 0.1) Medium (Ua = 0.4)

BMDb BMDratioc

BMDLb BMDLratioc

BMDb BMDratioc

BMDLb BMDLratioc

13 1233.5 1.001 1051.3 0.940 1237.6 1.005 866.3 0.77514 1039.1 0.999 790.0 0.876 1047.6 1.007 612.9 0.68015 1141.5 1.001 1030.9 0.964 1140.5 1.000 944.1 0.8837–9 687.7 1.001 631.1 0.971 687.3 1.000 586.3 0.90210–12 756.6 1.000 717.4 0.982 758.3 1.002 691.6 0.94713–15 693.8 1.003 607.2 0.952 697.0 1.007 530.2 0.8317–15 587.6 1.002 538.5 0.974 586.8 1.001 510.2 0.923

a Assumed intralitter correlation.b Log-Probit model, restricted to power P 1, BMR = 0.25 extra risk.c Transformed BMD or BMDL/untransformed BMD or BMDL, reference BBP untransfor

not shown), with differences as large as 6%. As with BBP, the differ-ences between the adjusted and unadjusted BMD ratios neither in-creased nor decreased with any particular pattern. Also similar toBBP, the differences between transformed and untransformedBMDLs were much larger than those seen with BMDs. Differencesbetween transformed and untransformed BMDLs were in somecases larger than 50% (i.e., the transformed BMDL was less than½ the untransformed BMDL), and a similar pattern was seen as thatwith BBP (i.e., the larger the assumed intralitter correlation, thelarger the difference between BMDLs). Because the differences be-tween BMDs were relatively small as with BBT, it was concludedthat using a model that did not take intralitter correlation into ac-count would make little difference to the outcomes with BMDsserving as the basis of comparison for TBT.

For BBP, the Hill model was chosen as the best fitting model forembryo lethality as a continuous variable; embryo lethality wasnot modeled as a continuous variable for TBT because it was notexplicitly reported as such in Ema et al. (1997). Acceptable fitswere obtained for data from all exposure regimens (using calcu-lated p-values and visual inspection as criteria for goodness-of-fit) (data not shown). Overall, the shapes of all the dose–responsecurves were very similar to those estimated by the log-Probit mod-el when embryo lethality was modeled as a dichotomous endpointas shown in Fig. 1.

Model estimates of BMD values at different selected BMRs (2.0,3.0, or 4.0 RD) for embryo lethality (as a continuous endpoint) werederived (data not shown). The BMDs calculated using a BMR of 2.0RD were most similar to those calculated using a BMR of 0.25 extrarisk when embryo lethality was modeled as a dichotomous end-point in Table 1. For example, the BMD calculated for exposureon GD 7–15 using continuous data and a BMR of 2.0 RD was594.8 mg/kg day as compared to 586.6 mg/kg day using the dichot-omized data in Table 1. BMDs for exposure on GD 13 (1155.1 mg/kg), 14 (1052.7 mg/kg), or 15 (1050.7 mg/kg) were on average 1.8times greater than the BMD for GD 7–15, compared to a ratio of1.9 when embryo lethality was modeled as a dichotomousendpoint.

3.3.2. Fetal weight reductionFor both male and female fetal weight loss, the Hill model was

the model choice, with Fig. 2 showing the modeled female dose–response curves for multiple-, intermediate-, and single-day expo-sures (male data not shown). TBT-induced fetal weight reductionwas not modeled due to non-monotonic dose–response curves,and incomplete reporting of data. Results in males and femaleswere qualitatively and quantitatively similar: exposure on GD 7–15 resulted in a sharp decrease in fetal weight with fetuses inthe highest dose group weighing approximately 20% less than

thalate, using Rao–Scott transformed, dichotomous fetal incidence data.

High (Ua = 0.7) Medium graded (Ua = 0.1, 0.3, 0.5, . . .)

BMDb BMDratioc

BMDLb BMDLratioc

BMDb BMDratioc

BMDLb BMDLratioc

1231.7 1.000 752.4 0.673 1235.3 1.003 910.6 0.8141024.0 0.984 521.6 0.579 1031.5 0.991 633.8 0.7031139.2 0.999 880.1 0.823 1146.4 1.005 984.6 0.921

689.7 1.004 540.6 0.832 700.4 1.020 611.6 0.941759.8 1.004 672.8 0.921 763.8 1.010 697.4 0.955690.3 0.998 463.8 0.727 705.6 1.020 581.4 0.911581.0 0.991 490.9 0.888 580.1 0.989 517.1 0.930

med BMD and BMDL values (Table 1), non-statistical comparison.

Fig. 2. Results from dose–response modeling of fetal weight in female rats for butylbenzyl phthalate (BBP). Observed data are denoted by symbols as labeled in the figure,closed symbols represent statistically significant dose–responses (p < 0.05 or 0.01) and open symbols represent non-significant dose–responses as reported by Ema et al.(1992, 1993, 1995, 1997, 1999) and Itami et al. (1990); lines represent model results. Data were fit with the Hill model. The modeled dose–response curves fit all theexperimental data acceptably using visual inspection as the criterion of goodness-of-fit although calculated p-values indicated poor fit, possibly as a result of pooling thecontrol values.

A. Davis et al. / Regulatory Toxicology and Pharmacology 54 (2009) 134–142 139

those in the control group. The intermediate-day exposures re-sulted in dose–response curves with similar shape but with verydifferent potencies. For example, exposure on GD 7–9 was mostsimilar to exposure on GD 7–15 with fetuses in the highest dosegroup also weighing approximately 20% less than control fetusesaccording to the model estimates. Exposures on GD 10–12 wereslightly less potent than those on GD 7–9 for both male and femalefetuses. Exposure on GD 13–15 was the least potent of the threeintermediate exposure regimens, with fetuses showing only a 5–7% decrease in body weight. Single exposures on GDs 13, 14, or15 did not result in any fetal weight reduction in either males orfemales despite administered doses that were nearly double thoseproducing the modest reductions noted for the 3-day exposures onGD 13–15.

The modeled BMD values at different selected BMRs for femalefetal weight reduction are shown in Table 3. BMDs at a BMR of 0.05RD calculated for exposure on GDs 7–9 were nearest that calcu-lated for GD 7–15 with ratios of 1.1–1.2. On the other hand, ratiosfor exposures on GD 10–12 and 13–15 were outside this range at1.4–1.5. BMDs were not calculated for GD 13, 14, 15 as there wasno reduction in fetal weight at any dose level compared to controlsfor these single-day exposures as described above.

3.3.3. Cleft palate and fused sternebraeModeling results for incidence of cleft palate and fused ster-

nebrae are presented in Fig. 3 and Table 4. Both abnormalities

Table 3BMD calculations for female fetal weight for butyl benzyl phthalate, using continuous litt

GD BMR = 0.05 RDa BMR = 0.10 RD

BMDb BMD ratio to GD 7–15 BMDb

13 >1500 >2.758 >150014 >1500 >2.758 >150015 >1500 >2.758 >15007–9 611.6 1.124 655.010–12 750.2 1.379 846.413–15 794.7 1.461 >10007–15 544.0 1.000 621.3

a RD, relative deviation.b Hill model, restricted to power P 1.

were modeled on the only appropriate data available, i.e., asdichotomous endpoints, for BBP exposures. Although there wasdata on tributyltin-induced increases in incidence for bothabnormalities, those data were of poor applicability for modelingpurposes and thus was not included in further analyses. For BBPand cleft palate, the Weibull model was the model of choice.Exposures on both GD 7–15 and 13–15 resulted in very steepdose–response curves, with both exposures resulting in 100%incidence of cleft palate. Exposures on GD 7–9 and 10–12 re-sulted in much lower incidences of cleft palate: 17% and 33%,respectively. Of the single-day exposures, only exposure on GD15 showed a substantial increase over background levels withan approximate incidence of 45%. The incidence on both GD 13and 14 was approximately 10%.

The BMD for cleft palate (at a BMR of 0.25 extra risk) calculatedfor exposure on GD 7–15 was 628.4 mg/kg. The BMD calculated forexposure on GD 13–15 (657.8 mg/kg) was essentially the same (ra-tio = 1.05) as the multiple-day BMD, whereas the BMD for GD 10–12 (925.3 mg/kg) was 1.47 times greater. The BMD calculated forexposure on GD 15 (1112.2 mg/kg) was nearly twice (1.77) theBMD for exposure on GD 7–15. BMDs were not calculated for GD13, 14, or 7–9 as there were insufficient dose–response informa-tion for any of those dosing regimens.

For BBP and fused sternebrae, the Weibull model was also themodel of choice. Exposures on both GD 7–15 and 13–15 resultedin very steep dose–response curves, with incidences of 57% and

er data.

a BMR = 0.15 RDa

BMD ratio to GD 7–15 BMDb BMD ratio to GD 7–15

>2.414 >1500 >2.140>2.414 >1500 >2.140>2.414 >1500 >2.140

1.054 712.3 1.0161.362 >1000 >1.426

>1.610 >1000 >1.4261.000 701.1 1.000

Fig. 3. Results from dose–response modeling of cleft palate and fused sternebrae for butylbenzyl phthalate (BBP); observed data are denoted by symbols as labeled in thefigure, closed symbols represent statistically significant dose–responses (p < 0.05 or 0.01) and open symbols represent non-significant dose–responses as reported by Emaet al. (1992, 1993, 1995, 1997, 1999) and Itami et al. (1990); lines represent model results. Cleft palate and fused sternebrae data were fit by the Weibull model. The modeleddose–response curves for cleft palate and fused sternebrae fit the experimental data acceptably for all exposures (using calculated p-values and visual inspection as criteriafor goodness-of-fit).

Table 4BMD calculations for malformations/variations for butyl benzyl phthalate, usingdichotomous litter incidence data.

GD Cleft palate Fused sternebrae

BMDa BMD ratio to GD 7–15 BMDa BMD ratio to GD 7–15

13 >1500 >2.387 >1500 >2.12414 >1500 >2.387 1487.1 2.10515 1112.2 1.770 893.7 1.2657–9 >1000 >1.591 >1000 >1.41610–12 925.3 1.472 >1000 >1.41613–15 657.8 1.047 657.8 0.9317–15 628.4 1.000 706.4 1.000

a Weibull model, restricted to power P 1, BMR = 0.25 extra risk.

140 A. Davis et al. / Regulatory Toxicology and Pharmacology 54 (2009) 134–142

100%, respectively. Exposures on GD 7–9 and 10–12 did not resultin any increase over background levels. Exposure on GD 15 re-

sulted in 100% incidence of fused sternebrae, whereas exposureson GD 13 resulted in no increase and on GD 14 an increase of29% increase over background. The BMD (at a BMR of 0.25 extrarisk) calculated for exposure on GD 7–15 was 706.4 mg/kg. TheBMD calculated for exposure on GD 13–15 (657.8 mg/kg) was verysimilar (1.07 times less) to the multiple-day BMD. The BMDs calcu-lated for exposure on GD 14 (1487.1 mg/kg) and GD 15 (893.7 mg/kg) were 2.11 and 1.27 times greater than the BMD for exposure onGD 7–15, respectively. BMDs were not calculated for GD 13, 7–9, or10–12, as there was no increase in incidence of fused sternebraefor any of those dosing regimens.

4. Discussion

In this study, we examined the occurrence of developmentaltoxicity results from laboratory animals for single-day exposures

A. Davis et al. / Regulatory Toxicology and Pharmacology 54 (2009) 134–142 141

as well as aspects and relationships between the single-day andthe corresponding multiple-day exposure scenarios.

The analysis of studies identified and collated in this currentwork gives clear indication that for a wide range of chemicalagents, developmental toxicity may result from an acute exposure.Further, the agents studied include many with an appreciable po-tential for human exposure (e.g., drugs, pesticides, and plasticiz-ers). For example, phthalic acid esters (including BBP) are widelyused as plasticizers for polyvinyl chloride products (Ema et al.,1999) such as floor tiles, medical devices, waterproof clothing,and food-wrapping products. The phthalic esters are not irrevers-ibly bound to the polymer and can migrate from the plastics intothe environment. Phthalates have been identified in marine andfreshwater environments, sediment, arable soils, and fish (Huet al., 2003; Hwang et al., 2006; Zeng et al., 2008; Fromme et al.,2002; Peijnenburg and Struijs, 2006). Estimates of human exposureto n-butyl phthalate, a mono-ester metabolite of BBP, have indi-cated that women of reproductive age appear to be exposed tohigher levels than the remainder of the population (Kohn et al.,2000).

Developmental effects occurring under single-day exposurescenarios do not appear to be qualitatively different or discernablefrom those occurring under multiple-day exposure scenarios. Thisobservation is anticipated and confirmatory for agents known toaffect specific endpoints within specific windows of susceptibilityoccurring during gestation (e.g., cleft palate on GD 14–16) (Thomasand Rossouw, 1991). However, this work indicates that this rela-tionship between multiple-day and single-day exposures may alsohave validity for endpoints, such as weight loss and fetal death,which are less specific and may not be associated with a particularor clearly distinguishable period in gestation.

Although limited in scope (only two chemicals were availablethat met data requirements), analyses from this work do give somepreliminary indication regarding the relationship of effective dosesbetween multiple-day exposure studies conducted throughoutgestation (i.e., typical guideline studies) and studies of exposureson single gestational days. The most general conclusion that canbe made from the data in this work is that higher doses are re-quired to elicit the same response level under a single-day expo-sure study than in the corresponding multiple-day exposurestudy. This observation extends to embryo lethality for both BBPand TBT, to fetal weight loss for BBP, and, marginally, to the highlyspecific ‘‘window-of-susceptibility” malformation of cleft palatefor BBP.

The contrasting half-lives of the toxicants included in the anal-ysis (BBT at 624 h and TBT at P24 h) allowed for extended analy-sis on the concept that with multiple exposures, the morepersistent compound (TBT) and its effects would accumulate muchmore so than would be anticipated with a less persistent com-pound (BBP). Indications of cumulative response in our analysiswould be an increased ratio between the equieffective responsedoses of the single-day/multiple-day paired studies (i.e., wherethe multiple-day appears increasingly potent relative to the sin-gle-day). It is noted that this ratio did increase for dichotomizedembryo lethality (Table 1), from around 2:1 for the less persistentcompound (BBP) to around 4:1 for the more persistent compound,TBT. This contrast with the less persistent agent, however, was notgreat (a factor of 2). Considerably more data would be required be-fore any definitive assertions could be made concerning thehypothesis that multiple-day studies of agents that are increas-ingly kinetically persistent would also exhibit increasing potencyrelative to a single-day exposure.

Although it is desirable to expand this type of analysis to moreexamples, several limitations of existing data for other chemicalsmay make this unfeasible at this time. Limitations encounteredin the current analytical portion of the work included low numbers

of experimental doses in the studies, inadequate qualitative andquantitative descriptions of the outcomes, inconsistent or inappro-priate dose–response characterizations, and the general lack ofindividual animal data necessary for the application of nestedmodels. Regarding the lack of individual animal data, this workalso demonstrates the use and value of the Rao–Scott transforma-tion in analysis of results of ‘‘nested” or, more generally, clustereddata. While we have shown that BMDLs do change considerablyaccording to the various assumed levels and patterns of intralittercorrelation, we cannot say which transformation most closelyapproximates the true intralitter correlation that exists in the data.This is why the central estimate (i.e., the BMD) was chosen as thebasis of comparison between dosing regimens. Transformed resultsfor BMD calculations were not distinguishable from the untrans-formed or original clustered data; these results provided a valuablemeans by which data utilization was maximized in this work. Fur-ther, this analysis demonstrates the need for the development ofguidance for modeling of developmental endpoints that ade-quately incorporates intralitter correlation when individual animaldata is lacking.

The goal of this work is to provide initial support for developingworking generalizations in the use of multiple-day developmentalstudies in acute health risk assessments. The data analyses in thiswork indicate that if an agent shows fetotoxicity in a guideline pre-natal developmental toxicity study, there is a strong potential forfetotoxicity under acute exposure conditions also. Further, the typeof endpoint reported in the guideline prenatal developmental tox-icity study is likely to be very similar if not the same under the con-ditions of an acute exposure. Also, limited information in this worksuggests that the dose–response information in the guideline studywould be paralleled by equieffective levels of response at higherexposure levels in the corresponding acute exposure study. Thissituation seemed to apply regardless of kinetics of the two agentsevaluated. In the examples here the magnitude of this differenceranged from near 1 to approximately fourfold lower for the guide-line study, depending on endpoint. However, detailed, quantitativedose–response analyses on a wider group of developmental toxi-cants would be necessary before proper generalizations in respectto the relationship between single- and multiple-day gestationalexposures would be appropriate.

Disclaimer

Partial funding was provided by the NCEA-ORISE InteragencyAgreement. The information in this document has been subjectedto review by the National Center for Environmental Assessment,U.S. Environmental Protection Agency, and approved for publica-tion. Approval does not signify that the contents reflect the viewsof the Agency, nor does mention of trade names or commercialproducts constitute endorsement or recommendation for use.The research presented in this document was funded in part bythe U.S. Environmental Protection Agency.

Conflict of interest

The authors declare that there are no conflicts of interest.

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