Derivation of Biomonitoring Equivalents for di-n-butyl phthalate (DBP), benzylbutyl phthalate (BzBP), and diethyl phthalate (DEP)
Introduction
Interpretation of measurements of concentrations of chemicals in samples of urine or blood from individuals in the general population is hampered by the general lack of screening criteria for evaluation of such biomonitoring data in a health risk context. Without such screening criteria, biomonitoring data can only be interpreted in terms of exposure trends, but cannot be used to evaluate which chemicals may be of concern in the context of current risk assessments. Such screening criteria would ideally be based on robust datasets relating potential adverse effects to biomarker concentrations in human populations (see, for example, the U.S. Centers for Disease Control and Prevention (CDC)1 blood lead level of concern; see http://www.cdc.gov/nceh/lead/). However, development of such epidemiologically-based screening criteria is a resource and time-intensive effort. As an interim approach, the development of Biomonitoring Equivalents (BEs) has been proposed, and guidelines for the derivation and communication of these values have been developed (Hays et al., 2007, Hays et al., 2008, LaKind et al., 2008).
A Biomonitoring Equivalent (BE) is defined as the concentration or range of concentrations of an environmental chemical (or metabolite) in a biological medium (blood, urine, or other medium) that is consistent with an existing health-based exposure guidance value such as a reference dose (RfD) or tolerable daily intake (TDI). Existing chemical-specific pharmacokinetic data are used to estimate biomarker concentrations that are consistent with the point of departure (POD) used in the derivation of an exposure guidance value (such as the RfD or TDI), and with the exposure guidance value itself. BEs can be estimated using available human or animal pharmacokinetic data (Hays et al., 2008), and BEs have been derived for numerous compounds including acrylamide, cadmium, 2,4-dichlorophenoxyacetic acid, toluene, and others (reviewed in Hays and Aylward (2009)). BEs are intended to be used as screening tools to allow an assessment of biomonitoring data to evaluate which chemicals have large, small, or no margins of safety compared to existing risk assessments and exposure guidance values. BE values are only as robust as are the underlying exposure guidance values and pharmacokinetic data used to derive the values. BE values are not intended to be diagnostic for potential health effects, but rather are risk management tools for use in evaluating biomonitoring data in the context of existing chemical risk assessments.
This manuscript presents derivation of BE values for three diester phthalate compounds: diethyl phthalate (DEP; Chemical Abstracts Services [CAS] Registry number 84-66-2), benzylbutyl phthalate (BzBP; CAS number 85-68-7), and di-n-butyl phthalate (DBP; CAS number 84-74-2) (Fig. 1). The metabolites of these three compounds are included in the list of analytes that will be measured in urine in the upcoming Canadian Health Measures Survey, thus of current interest in terms of BE derivation. A companion manuscript describes the BE derivation for di-2(ethylhexyl) phthalate (DEHP). The evaluation of these three phthalate compounds is presented separately because the DEHP BE derivation accounts for multiple metabolites (oxidative metabolites in addition to mono-ester metabolites) and there is a more detailed literature available on DEHP.
The three phthalates included in this manuscript, and other diester phthalate compounds, are used for a wide range of applications in consumer products including in cosmetics and personal care products and as plasticizers in the production of plastics for a wide range of applications including food packaging materials, enteric coatings for pharmaceuticals, and in medical devices (NRC, 2008). The routes and sources of exposure to phthalates for persons in the general population vary by phthalate compound. DBP is used as a coalescing aid in latex adhesive formulations, as a plasticizer for cellulosic plastics, and as a solvent for dyes (Kavlock et al., 2002a). BzBP is primarily used as a plasticizer in the production of polyvinyl chloride and other plastics, which are then used in a variety of sealing, coating, painting, and adhesive products and formulations, Kavlock et al., 2002b, EFSA, 2005b, EFSA, 2005a, Wormuth et al., 2006, Heudorf et al., 2007). DEP is used as a vehicle for fragrances, cosmetics, and personal care products (Api, 2001, Wormuth et al., 2006). In an evaluation of the routes and sources of exposure for European population, Wormuth et al. (2006) concluded that the principle route of exposure for DBP for persons in the general population was through trace levels in foods, followed by some exposure via inhalation pathways. Exposure to BzBP was attributed to foods, spray paints, and in small children, ingestion of dusts or through mouthing behaviors on plastic objects. Exposure to DEP was estimated to result mainly from dermal absorption of DEP from personal care products, followed by inhalation exposure.
Phthalates cause a variety of toxic effects in laboratory animals including effects on the liver and reproductive system. However, effects on the reproductive system of male laboratory rats exposed in utero and during development have emerged as endpoints of particular concern for several phthalate compounds. Evaluations of phthalates with a range of structures has demonstrated that those with side chains four to six carbons in length in the ortho positions of the molecule displayed specific toxicity to the developing male reproductive tract, while those with shorter branches generally did not, although this pattern is not absolute (NRC, 2008). Consistent with this structure–activity relationship, of the three phthalates included in this evaluation, DEP has consistently been inactive in assays designed to detect adverse effects on the developing male rat reproductive system, while BzBP and DBP have caused reduced fetal testosterone and other effects indicating antiandrogenic activity (Howdeshell et al., 2008a, NRC, 2008). Metabolites of phthalate compounds have been measured in urine in numerous population biomonitoring studies including those conducted by the CDC (2005) and the German Human Biomonitoring Commission (http://www.umweltbundesamt.de/gesundheit-e/monitor/index.htm) as well as in targeted studies designed to assess exposures to and potential human health effects of phthalate compounds.
Section snippets
Available data and approach
Exposure guidance values, critical effects, and mode of action. Exposure guidance values from national and international agencies were reviewed and identified. The focus of this review was on values derived by the USEPA, U.S. ATSDR, Health Canada, and agencies of the European Union (EFSA and the ECB). Table 1 presents the available chronic exposure guidance values derived for DEP, DBP, and BzBP. For each guidance value, the point of departure (POD), the toxicological endpoint of interest, and
Available pharmacokinetic data—laboratory animals
The pharmacokinetics of DBP have been investigated in rats in experimental studies (NIEHS, 1995, Fennell et al., 2004, Clewell et al., 2009), and a comprehensive PBPK model for DBP in pregnant rats has been developed (Clewell et al., 2008). The model and available datasets allow estimation of a variety of biomarker concentrations in rats, including maternal plasma concentration of MBP, at dose levels in the range of the doses administered in the critical study underlying the USEPA RfD for DBP
BE derivation
The selection of mono-ester metabolites in urine as the most reliable biomarkers for these three phthalate compounds dictates a mass balance approach with an assumption of steady-state intake and excretion. Specifically, the amount of mono-ester metabolite excreted in urine each day will be approximately equal to the amount ingested at the TDI or RfD times a factor to account for the excretion fraction for each specific mono-ester metabolite (see Table 2).
The process of BE derivation for each
Sources of variability and uncertainty
Several sources of variability and uncertainty are associated with the BE values presented in Table 5. One source of variability that will impact the measured concentrations in urine is the relatively short half-life of excretion of the mono-ester metabolites. Anderson et al. (2001) found no detectable labelled metabolite from DEP or DBP eliminated after the first 24 h. Specific estimates of urinary half-life were not calculated; however, the essentially complete elimination of labelled compound
Confidence assessment
The guidelines for derivation of BE values (Hays et al., 2008) specify consideration of two main elements in the assessment of confidence in the derived BE values: robustness of the available pharmacokinetic data and models, and understanding of the relationship between the measured biomarker and the critical or relevant target tissue dose metric. As discussed above, the pharmacokinetic data for DBP and BzBP were derived from a study of excretion fraction in 16 individuals at two dose levels.
Discussion and interpretation of BE values
The BE values presented here represent estimates of the 24-h average concentrations of mono-ester metabolites in urine that are consistent with the existing exposure guidance values for DBP, DEP and BzBP resulting from the risk assessments conducted by various governmental agencies as listed in Table 1. The values were derived based on current understanding of the pharmacokinetic properties of these compounds in humans. No BE values based on serum concentrations were derived at this time due to
Conflict of interest satement
The authors declare they have no conflicts of interest.
Acknowledgments
Funding for this project was provided under Health Canada Contract 4500195930. The views expressed in this article are those of the authors and do not necessarily reflect the views or policies of Health Canada. We acknowledge Risk Sciences International for conducting an independent peer-review to assure the BE derivations presented here are consistent with the guidelines for the derivation of BEs (Hays et al., 2008) and that the best available science, data and/or models were used to calculate
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