Derivation of Biomonitoring Equivalents for di-n-butyl phthalate (DBP), benzylbutyl phthalate (BzBP), and diethyl phthalate (DEP)

Derivation of Biomonitoring Equivalents for di-n-butyl phthalate (DBP), benzylbutyl phthalate (BzBP), and diethyl phthalate (DEP)

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Regulatory Toxicology and Pharmacology 55 (2009) 259–267

Contents lists available at ScienceDirect

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

Derivation of Biomonitoring Equivalents for di-n-butyl phthalate (DBP), benzylbutyl phthalate (BzBP), and diethyl phthalate (DEP) Lesa L. Aylward a,*, Sean M. Hays b, Michelle Gagné c, Kannan Krishnan c a

Summit Toxicology, LLP, 6343 Carolyn Drive, Falls Church, VA 22044, USA Summit Toxicology, LLP, Lyons, CO, USA c Université de Montréal, Département de santé environnementale et santé au travail, Montréal, QC, Canada b

a r t i c l e

i n f o

Article history: Received 25 August 2009 Available online 12 September 2009 Keywords: Biomonitoring Risk assessment Phthalates Pharmacokinetics

a b s t r a c t Recent efforts worldwide have resulted in a growing database of measured concentrations of chemicals in blood and urine samples taken from the general population. However, few tools exist to assist in the interpretation of the measured values in a health risk context. Biomonitoring Equivalents (BEs) are defined as the concentration or range of concentrations of a chemical or its metabolite in a biological medium (blood, urine, or other medium) that is consistent with an existing health-based exposure guideline, and are derived by integrating available data on pharmacokinetics with existing chemical risk assessments. This study reviews available health-based exposure guidance values for di-n-butyl phthalate (DBP), benzylbutyl phthalate (BzBP), and diethyl phthalate (DEP) from Health Canada, the United States Environmental Protection Agency (U.S. EPA), the Agency for Toxic Substances and Disease Registry (ATSDR), and the European Food Safety Authority (EFSA). BE values corresponding to the oral reference dose (RfD), minimal risk level (MRL) or tolerable daily intake (TDI) estimates from these agencies were derived for each compound based on data on excretion fractions of key urinary metabolites. These values may be used as screening tools for evaluation of biomonitoring data for metabolites of these three phthalate compounds in the context of existing risk assessments and for prioritization of the potential need for additional risk assessment efforts for each of these compounds relative to other chemicals. Ó 2009 Elsevier Inc. All rights reserved.

1. 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 * Corresponding author. E-mail address: [email protected] (L.L. Aylward). 1 Abbreviations used: ADI, acceptable daily intake; AR, androgen receptor; ATSDR, Agency for Toxic Substances and Disease Registry; BE, Biomonitoring Equivalent; BEPOD, Biomonitoring Equivalent Point of Departure; BMDL, benchmark dose lower limit; BzBP, benzylbutyl phthalate; DBP, di-n-butyl phthalate; DEP, diethyl phthalate; DEHP, di(2-ethylhexyl) phthalate; DiBP, diisobutyl phthalate; DMP, dimethyl phthalate; EFSA, European Food Safety Authority; MBP, monobutyl phthalate; MBzP, mono-benzyl phthalate; MEP, mono-ethyl phthalate; MW, molecular weight; NOAEL, No Observed Adverse Effect Level; POD, point of departure; RfD, reference dose; USEPA, United States Environmental Protection Agency; UF, uncertainty factor; TDI, tolerable daily intake. 0273-2300/$ - see front matter Ó 2009 Elsevier Inc. All rights reserved. doi:10.1016/j.yrtph.2009.09.003

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, 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.

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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, 2005a,b; 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.

2. 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 the applied uncertainty factors are summarized. The TDI values derived for DBP and BzBP by the European Food Safety Authority (EFSA, 2005a,b) are the most recent exposure guidance values for these compounds. These TDIs are both based explicitly on developmental effects on the reproductive system of male rodent offspring exposed in utero and during development. Endpoints falling into this category have emerged as the endpoints of greatest concern in current risk evaluations for selected phthalate compounds, including DBP and BzBP (Howdeshell et al., 2008a; NRC, 2008). In contrast, evaluations of potential DEP toxicity to the developing male rat reproductive system have consistently found no effects at doses equal to or above the NOAEL of 750 mg/kg/d based on other endpoints including decreased growth rate and altered organ weights (Gray et al., 2000; Howdeshell et al., 2008b). The modes of action for selected phthalates in producing adverse effects on the developing male reproductive system are under investigation. However, several specific mechanisms have been investigated, including alteration of fetal testosterone metabolism leading to reductions in testosterone concentrations, effects on germ cell formation and maturation, and effects on the expression of the insl3 gene, which is important for normal testicular descent (data reviewed in NRC (2008)). The issue of whether parent

Fig. 1. Structures of (a) diethyl phthalate, molecular weight 222; (b) benzylbutyl phthalate, molecular weight 312; (c) di-n-butyl phthalate, molecular weight 278.

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L.L. Aylward et al. / Regulatory Toxicology and Pharmacology 55 (2009) 259–267 Table 1 Current chronic exposure guidance values from Health Canada, the US Environmental Protection Agency (USEPA), and the European Food Safety Authority. Parent Compound

Reference Value mg/kg/d

DBP

TDI, Health Canada (1994)

6.3  102

TDI, EFSA (2005a)

1  102

RfD, USEPA (1990)

1  101

TDI, Health Canada (2000)

1.3

TDI, EFSA (2005b)

5  101

RfD, USEPA (1993a)

2  101

RfD, USEPA (1993b)

8  101

BzBP

DEP

Details  NOAEL: 62.5 mg/kg/d  UF: 10 (interspecies variation); 10 (intraspecies sensitivity); 10 (severity of effects)  Critical effects: Decreases in live offspring and increases in external defects and skeletal anomalies in mice  LOAEL: 2 mg/kg/d  UF: 200 (attribution of components not discussed)  Critical effects: Loss of germ cell development and mammary gland changes in rats exposed via diet during gestation through lactation  NOAEL:125 mg/kg/d  UF: 10 (interspecies variation); 10 (intraspecies sensitivity); 10 (subchronic to chronic + study deficiencies)  Critical effects: Increased mortality in rats given DBP in diet for 1 year  BMDL05: 132 mg/kg/d  UF: 10 (interspecies variation); 10 (intraspecies sensitivity)  Critical effects: Increases in pancreatic lesions in male rats exposed via dietary administration for 13 weeks  NOAEL: 50 mg/kg/d  UF: 100 (attribution of components not discussed)  Critical effects: reduction in anogenital distance in F1 and F2 rat offspring in a dietary administration multi-generational study  NOAEL: 159 mg/kg/d  UF: 10 (interspecies variation); 10 (for intraspecies sensitivity); 10 (subchronic to chronic)  Critical effects: Increased liver:body and liver: brain weight ratios in rats administered BzBP in the diet for 26 weeks  NOAEL: 750 mg/kg/d  UF: 10 (interspecies variation); 10 (intraspecies sensitivity); 10 (subchronic to chronic)  Critical effects: Decreased growth rate, food consumption and altered organ weights in rats given DEP in diet for 16 weeks

compounds or metabolites produce these effects has not been fully investigated; however, in general, evidence appears to suggest that metabolites of the phthalate compounds are primarily responsible for the observed toxicities (reviewed in Kavlock et al. (2002a,b) and Clewell et al. (2009)). Available pharmacokinetic data—humans. We identified key studies describing the pharmacokinetics of these three compounds based on Medline searches and recent literature reviews (e.g., NRC, 2008). We focused on identification of data sets that allow quantitative evaluation of metabolites in human urine in relationship to a known administered dose. Experimental data in humans describing the urinary elimination of metabolites of DBP and BzBP following oral administration are summarized in Table 2 (Anderson et al., 2001). 13C-labelled parent compounds were administered to groups of 8 individuals at a low and high dose for each compound, and 24-h urine samples were collected on the day of administration, the next day, and day 6. The DBP doses used were 255 and 510 lg; assuming that the volunteers had bodyweights of approximately 70 kg (actual bodyweights not reported) these doses correspond to 3.6 and 7.3 lg/kg-d. The BzBP doses used were 253 and 506 lg (3.6 and 7.3 lg/kg-d). These experimental doses are substantially lower than the various reference doses and tolerable daily intake estimates for these compounds (Table 1). No labelled phthalate derivatives were found in samples collected after the first day, confirming the rapid metabolism and

elimination of these compounds in urine, and confirming that the excretion fractions measured account for the full administered doses (Anderson et al., 2001). The analytical methods employed in this study captured both free and glucuronide conjugated mono-ester metabolites. There was no difference in molar fraction excreted as the phthalate mono-ester between the high and low dose for either compound. The mass excretion fractions were estimated by multiplying the molar excretion fractions by the ratio of the molecular weights of the metabolites and parent compounds. Several studies suggest that in both humans and rats, dialkyl phthalates are metabolized more or less completely in the intestinal tract and absorption occurs as the mono-ester metabolites (Lake et al., 1977; Rowland et al., 1977). The elimination of 70 to 80% of the administered dose as the mono-ester in urine (rather than 100%) suggests that some additional oxidative metabolism occurs following absorption in humans (Calafat and McKee, 2006; Silva et al., 2007;). MBP was the principal urinary metabolite of DBP, but it was also observed as a minor metabolite of BzBP; see Table 2 (Anderson et al., 2001; Silva et al., 2007). Because MBP can be produced both as the major metabolite of DBP and as a minor metabolite of BzBP, urinary excretion of this metabolite should be used as a marker of DBP exposure with some caution. If substantial amounts of MBzP are present in a urine sample from an individual (suggesting exposure to BzBP), the amount of MBP potentially attributable to BzBP

Table 2 Excretion fraction of phthalate metabolites in human urine. Phthalate (MW)

Metabolite (MW)

Molar excretion fraction (%)

Mass excretion fraction (%)

Reference

DBP (278) BzBP (312)

Mono-butyl phthalate (MBP) (222) Mono-benzyl phthalate (MBzP) (256) Mono-butyl phthalate (MBP) (222) Mono-ethyl phthalate (MEP) (194)

69 73 6 70

55 60 5 61

Anderson et al. (2001)a Anderson et al. (2001)a Anderson et al. (2001)a –

DEPb (222) a

Average of measurements of total excretion of labelled doses from two groups of 8 individuals. Relative standard deviations averaged approximately 30%. No experimentally derived excretion factor available for DEP in humans. Molar excretion fraction estimated was set to 70% MEP based on similarity to DBP (Calafat and McKee, 2006) and the corresponding mass excretion fraction was estimated based on the ratio of molecular weights of MEP and DEP. b

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should be considered in the evaluation of the MBP concentration as a marker of DBP exposure. Hydroxylated and carboxylated metabolites of DBP are also present, but account for a minor fraction of the total dose (Silva et al., 2007). The metabolism of BzBP occurs preferentially to mono-benzyl phthalate (MBzP). Thus, the quantification of MBzP provides a relatively robust method for quantification of BzBP exposure (Anderson et al., 2001). DEP was not evaluated in the study by Anderson et al. (2001). However, based on data from animal studies, metabolism of DEP has been postulated to occur similarly to metabolism of DBP, with a similar excretion fraction as the mono-ester, mono-ethyl phthalate (MEP), expected (Lake et al., 1977; Koch et al., 2003; Calafat and McKee, 2006). Urinary excretion of metabolites of DEP and DBP was also observed following dermal application in adult male volunteers (Janjua et al., 2008), demonstrating the potential for dermal absorption of these compounds. Peak urinary concentrations occurred approximately 8–12 h following topical application. Serum concentration data in humans following controlled exposures were not identified for these parent compounds or their metabolites. 3. 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 and other studies of interest. No data on the pharmacokinetics of DBP in mice (the species used in the derivation of the Health Canada TDI) were identified. Limited pharmacokinetic data are available for BzBP in dogs and rats, primarily demonstrating the potential for dermal absorption and the dominance of urinary elimination of the two mono-ester metabolites (MBzP and MBP) at lower exposure levels (reviewed in Kavlock et al. (2002b)). Basic pharmacokinetic data were developed for DEP in rats and mice following dermal and oral administration (reviewed in Api (2001)). These data generally confirm MEP as the major metabolite with urinary excretion providing the major route of elimination. Potential biomarkers. The potential biomarkers for DBP, BzBP, and DEP are summarized in Table 3. There is evidence that the ma-

jor toxicological endpoints observed after administration of the parent phthalate compounds reviewed here are due to the monoester metabolites (reviewed in Kavlock et al. (2002a,b)). This suggests that the concentration of free mono-ester metabolites in blood or serum could be relevant internal dose metrics for risk assessment and thus valuable biomarkers of exposure. However, analysis of phthalates or their mono-ester metabolites in serum or whole blood is complicated by several factors including the widespread presence of the phthalate compounds in laboratory and sample collection equipment (leading to the possibility of contamination of samples) and the presence of esterase activity in serum (Blount et al., 2000). This esterase activity can result in ongoing metabolism of phthalate compounds to mono-ester derivatives in serum following sample collection, even under conditions of low temperature storage (Kato et al., 2003; Koch et al., 2005). Finally, the rapid metabolism of these compounds results in very short half-lives in serum (estimated to be under 2 h), complicating the design and interpretation of studies of these phthalates in serum or whole blood in relation to external exposures (Anderson et al., 2001; Koch et al., 2005). Urinary half-lives of the mono-ester metabolites are also substantially less than a day; however, the kinetics of urinary excretion and elimination allow a somewhat longer window for collection and analysis of metabolites from phthalate exposure than afforded by serum or whole blood. In addition, these mono-ester metabolites in urine are not subject to the possibility of ongoing metabolism after sample collection as may occur in serum samples. Thus, the mono-ester metabolites of these parent compounds in urine are currently the most reliable biomarkers available.

4. 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 of these compounds is illustrated in Fig. 2 and described as follows: 1. Identify the POD used as the basis for the derivation of the TDI or RfD.

Table 3 Potential biomarkers of exposure to DEP, DBP and BBzP.

Parent compound

Primary metabolite

Analyte

Medium

Advantages

Disadvantages

DEP DBP BzBP DEP DBP BzBP

Serum



Short half-life; Invasive

Urine



Not found in urine

Serum



Short half-life; invasive

Urine

Appears to be the major urinary metabolite for each of the phthalates; specific biomarkers of exposure*; non-invasive

Lack of detailed kinetic information on each of the mono-esters as well as their downstream metabolites

MEP MBP MBzP MEP

MBP

MBP is produced as a minor metabolite of BzBP, and so interpretation as a marker of DBP exposure only must be made cautiously

MBzP *

MBP is not a specific biomarker for DBP because it is also produced as a minor metabolite of BzBP, however, the relative amount of MBP produced from BzBP is quite low compared to production from DBP.

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Relevant Internal Dose

External Dose

Human Equiv. POD

Estimate metabolite Estimate metabolite output outputusing using excretion fraction excretion fraction data; data;divide divideby byavg. avg.daily daily creatinine excretion creatinine excretion or orurinary urinaryvolume volume

BEPOD

BE

Fig. 2. General schematic for derivation of urinary BE values for mono-ester metabolites of DEP, BzBP, and DBP. *Animal POD includes adjustments for duration, and severity of effect. UFA: Interspecies uncertainty factor; UFH: Intraspecies uncertainty factor.

2. Apply any uncertainty factors (UFs) used in the TDI or RfD derivation to account for exposure duration, severity of endpoint, and interspecies extrapolation to identify the human-equivalent POD. 3. Estimate the total daily urinary excretion of mono-ester on a molar and mass basis per unit dose of parent compound. Divide the estimated daily excretion per unit dose of parent compound by 24-h average urine volume and average creatinine excretion. This factor allows estimation of average urinary concentrations associated with steady-state chronic intake of any dose of parent compound (Table 4). 4. Apply the urinary excretion factor to the human-equivalent POD from step 2 to estimate the urinary BEPOD. 5. Apply the intraspecies uncertainty factor to the BEPOD to derive the BE. Specifically, the estimated urinary concentration on a volume basis for each metabolite associated with a unit dose of parent

D  BW  UEF V

ð1Þ

where CV is the average urinary concentration on a volume basis of the metabolite, D is a unit dose of parent compound (1 lg/kg-d), BW is the bodyweight for the group, UEF is the mass-based urinary excretion fraction for the metabolite, and V is the 24-h average urinary volume. Similarly, the creatinine-adjusted concentration was calculated as follows:

CC ¼

UFH

Human

compound for a specific population sub-group was calculated using the following formula:

CV ¼

Animal POD* UFA

Animal

Monitored Biomarker

D  BW  UEF CE

ð2Þ

where CC is the creatinine-adjusted 24-h urinary concentration of the metabolite, and CE is the 24-h creatinine excretion rate. Data on urinary volume and creatinine excretion rates were drawn from a variety of studies noted in footnotes to Table 4. Because for each key metabolite the average metabolite concentrations associated with a unit dose of parent phthalate varied little across age groups, the average across all age groups was estimated and carried forward in the calculations. This approach also allows a direct translation of alternative dose levels of interest into expected average urinary concentrations for each of the phthalate compounds. For example, if new assessments of TDI or RfD values are issued, the factors expressing urinary concentrations as a function of unit dose of parent compounds presented in Table 4 allow direct calculation of corresponding BE values. Using the estimates of average urinary concentrations of metabolite per unit dose of parent compound, estimates of the average urinary concentrations associated with each of the exposure guidance values from Table 1 were estimated at the human-equivalent POD and at the exposure guidance value (Table 5). 5. 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

Table 4 Estimated steady-state urinary concentrations of mono-ester phthalate metabolites per unit dose of parent compound by age group. Age group

a

Bodyweight kga

Children, 6–11

32

Adolescents, 11–16

57

Men, >16

70

Women, >16

55

Average 24 h urinary volume, L (creatinine excretion, g)b,c

24 h average urinary concentration per ug/kg/d administered dose of parent, lg/L (lg/g creatinine) MBPd

MBzP

MEP

0.66 (0.50) 1.65 (1.20) 1.70 (1.50) 1.60 (1.20)

26.7 (35.3) 19.0 (26.2) 22.7 (25.7) 18.9 (25.3)

29.0 (38.3) 20.7 (28.5) 24.7 (28.0) 20.6 (27.5)

29.7 (39.1) 21.1 (29.1) 25.2 (28.5) 21.0 (28.0)

Average, lg/L: (lg/g creatinine):

21.8 (28.1)

23.7 (30.5)

24.3 (31.2)

Average bodyweight, estimated from Table 8–1 of USEPA (2008). Average 24-h urinary volumes for children from Remer et al. (2006). Volumes for adults from Perucca et al. (2007). Adolescents were assumed to have urinary volumes similar to average values for adults. c Average 24-h creatinine excretion for children and adolescents estimated from Remer et al. (2002); average creatinine excretion for boys and girls under age 13, 17 mg/kg BW per day; average creatinine excretion for adolescents, 22 mg/kg BW per day. Creatinine excretion for adults estimated based on equations from Mage et al. (2004) and average US height and specified bodyweights. d MBP is estimated here based only on DBP as parent. The minor contribution from BzBP is not quantified here. b

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Table 5 Derivation of BE values consistent with exposure guidance values from Table 1. The POD values and uncertainty factors (UFs) are described in more detail in Table 1. Parent compound and monitored analyte: Exposure guidance value:

DEP as MEP

BzBP as MBzP

DBP as MBP

RfD, USEPA (1993b)

TDI, Health Canada (2000)

TDI, EFSA (2005b)

RfD, USEPA (1993a)

TDI, Health Canada (1994)

TDI, EFSA (2005a)

RfD, USEPA (1990)

POD, mg/kg-d: UFL, UFS, UFA: Human-equivalent POD, mg/kg-d: BEPOD, mg/L in urine: (mg/g creatinine):

750 100 7.5

132 10 13.2

50 10 5

159 100 1.59

62.5 100 0.625

2 20 0.1

125 100 1.25

180 (230)

310 (400)

120 (150)

38 (49)

14 (18)

2 (2.8)

27 (35)

UFH: BE, mg/L in urine: (mg/g creatinine):

10 18 (23)

10 31 (40)

10 12 (15)

10 3.8 (4.9)

10 1.4 (1.8)

10 0.2 (0.28)

10 2.7 (3.5)

of urinary half-life were not calculated; however, the essentially complete elimination of labelled compound within 24 h suggests half-lives of less than 6 h, and possibly much shorter. With elimination half-lives on this order, urinary concentrations are likely to vary substantially over the course of 24 h. Such potential variations were examined using a simple pharmacokinetic model for DEHP (Lorber et al., in press) which has metabolites with similar urinary excretion half-lives (Aylward et al., 2009). These simulations demonstrated that for a given daily exposure rate, under a reasonable range of assumptions regarding timing of exposures and urinary spot sample collection, spot sample urinary concentrations could be expected to exhibit concentrations 2 to as much as 5 times higher than the 24 h average urinary concentrations. Likewise, if urinary spot samples are collected many hours after last exposure (for example, in a first morning void sample) concentrations may be substantially lower than the 24 h average concentration. These results are also consistent with the limited available data on variations within individuals in spot urine sample concentrations over the period of several days (see Preau et al. (2007), for example). For this reason, the guidelines for BE derivation (Hays et al., 2008) specify that, for compounds with short biological half-lives, comparisons between BE values and population biomonitoring data should be made on a population basis using the central tendency of the population values rather than the extremes. This analysis also suggests that more reliable evaluations of exposure levels could be made based on 24-h urine collections rather than spot samples. Other sources of potential variation in measured urinary concentrations, even under conditions of exposure consistent with the RfD, include variations in hydration status and creatinine excretion rates, and individual variations in metabolism patterns, which could impact measured concentrations in a spot urine sample by a factor of two to three (Scher et al., 2007). The appropriateness of adjustment for hydration status using creatinine excretion has been debated (Garde et al., 2004; Barr et al., 2005; Scher et al., 2007) because creatinine excretion also can vary substantially due to variations in dietary pattern as well as other individual factors, and the variability in children may be greater than in adults (O’Rourke et al., 2000; Kissel et al., 2005). However, we have no information to suggest that creatinine adjustment for measures of the mono-ester phthalate metabolites in urine is particularly inappropriate. Thus, we provide this additional calculation of BE values on that basis as well as on a urinary volume basis for use when results are reported on a creatinine-adjusted basis. Further discussion of issues surrounding creatinine adjustment is presented in Hays and Aylward (2009). Samples collected for a 24-h period would be expected to be less influenced than spot samples by both variations in hydration status and creatinine excretion. Other sources of uncertainty affect the interpretation of the BE values presented here. The pharmacokinetic data on urinary excre-

tion fraction for the mono-ester metabolites of DEP and DBP come from a single study of controlled exposure with eight adult volunteers in each of two dose groups. The degree of variability among individuals in the study was relatively modest, with relative standard deviations of approximately 30%, however, the experimental population likely does not represent the full range of variability that may occur in the general population. Furthermore, no chemical-specific data for DEP was collected in Anderson et al. (2001) or any other study; the assumption has been made here and elsewhere (Koch et al., 2003; Calafat and McKee, 2006) that the excretion fraction as MEP is similar as observed for the other mono-ester phthalates. Overall, no data are available to assess the degree of likely inter-individual variability due to age, sex, genetic polymorphisms, or to modifying factors such as smoking, alcohol consumption or other factors. Thus, the potential variation due to these factors is unknown. One source of uncertainty in the application of these BE values to interpretation of biomonitoring data in the general population is the co-occurrence of MBP as a metabolite of both DBP and BzBP. MBP is the major metabolite of DBP; however, it also is a minor metabolite of BzBP. BzBP is cleaved preferentially to produce MBzP, but Anderson et al. (2001) found that in the higher dose group, approximately 6% of BzBP was converted to MBP. Under conditions in which levels of exposure to BzBP are not dramatically higher than to DBP, the vast majority of MBP detected can be interpreted as occurring due to exposure to DBP. However, if exposure to BzBP is substantially higher than to DBP, the minor production of MBP may confound the attribution of the source of the observed MBP in urine samples. Interpretation of observed MBP should take into account the co-observed levels of MBzP as well, in order to assess this possibility. Finally, the analytical specificity of the analysis of MBP in a given study should be assessed. The BE values presented here are based on both toxicity and pharmacokinetic data for di-n-butyl phthalate. However, some analytical methods for mono-butyl phthalate in urine do not separate the mono-n-butyl phthalate from mono-isobutyl phthalate (the mono-ester metabolite of diisobutyl phthalate), including both in the reported totals. Since di-isobutyl phthalate is a separate compound used, in some cases, in distinct applications, and which may potentially have a different degree of toxicity or different excretion fraction than di-n-butyl phthalate (Heudorf et al., 2007; NRC, 2008), interpretation of urinary concentration data which does not resolve these two isomers must be made with care.

6. Confidence assessment The guidelines for derivation of BE values (Hays et al., 2008) specify consideration of two main elements in the assessment of

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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. However, no similar study is available for DEP. Because the urinary mass balance data suggest excretion of the majority of compound for DBP and BzBP as the simple mono-ester metabolite, the estimates presented here should be relatively reliable as a screening tool, although the full range of variability potentially present in the population has probably not been represented in this dataset. Another potential limitation of the dataset is the dose range used in the collection of the data. The doses used in Anderson et al. (2001) were approximately 3 and 7 lg/kg-d for DBP and BzBP, while the chronic exposure guidance values for these compounds are substantially higher (10 to 100 lg/kg-d for DBP; 200 to 1300 lg/kg-d for BzBP; see Table 1 for details). Thus, use of these data to estimate urinary excretion levels at the human-equivalent POD or the exposure guidance values requires extrapolation from the range of the observed data. We have no information to suggest that metabolism patterns would be substantially different at the higher exposure levels, and so cannot address this potential uncertainty further. The urinary biomarker concentrations used here are not directly related to target tissue concentrations of the relevant toxic moieties. However, the urinary excretion rates of the metabolites are likely to be directly related to serum concentrations of the various metabolites, which are believed to be the toxicologically active moieties. Thus, the assessment of the confidence level in the derived BE values based on these two factors is as follows:  Robustness of pharmacokinetic data: MEDIUM (for BzBP and DBP); LOW (for DEP).  Relevance of biomarker to relevant dose metrics: MEDIUM.

7. 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 inadequacies in the database, the very short serum half-lives of the parent and metabolite compounds, and due to the fact that current biomonitoring efforts for these compounds are directed at urinary metabolites. However, if the database improves and/or biomonitoring for these compounds or their metabolites in serum is contemplated, this decision could be revisited based on the limited available data. These BE values should be regarded as interim screening values that can be updated or replaced if the exposure guidance values are updated or if the scientific and regulatory communities develop additional data on acceptable or tolerable concentrations in human biological media. The appropriate uses and limitations of BE values, particularly for compounds with relatively short biological and urinary halflives such as the phthalates considered here and their metabolites, have been discussed previously (Hays et al., 2008; Hays and Aylward, 2009). The BE values can be used as a screening tool to evaluate population- or cohort-based biomonitoring data in the context of existing risk assessments. As discussed above, because of the likely variations in measured urinary concentrations result-

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ing from the relatively short biological half-lives of these compounds, the BE values are best used for assessment of the central tendency of measured values in a study. In accordance with the guidelines for communication using BE values (LaKind et al., 2008), measured urinary concentrations below the BE values represent relatively low priority for risk assessment follow-up. Concentrations in excess of the BE values, but less than the BEPOD values represent medium priority for risk assessment follow-up, while those in excess of the BEPOD indicate high priority for risk assessment follow-up. Based on the results of such comparisons, an evaluation can be made of the need for additional studies on exposure pathways, potential health effects, other aspects affecting exposure or risk, or other risk management activities. BE values do not represent diagnostic criteria and cannot be used to evaluate the likelihood of an adverse health effect in an individual. Measured values in excess of the identified BE values may indicate exposures at or above the current exposure guidance values that are the basis of the BE derivations. However, as discussed above, measured concentrations above the BE values, which are based on 24-h average urinary concentrations, would be expected even if exposures do not exceed the exposure guidance values due to the transient concentration profiles in urine expected for these compounds, variations in hydration status, and other factors discussed further above. Thus, interpretation of data for individuals or of tails of the distribution in population-monitoring studies is not appropriate. In addition, the exposure guidance values for these compounds were derived with a substantial margin from doses that resulted in no observed effect in the most sensitive animal toxicity studies. Thus, these values are not ‘‘bright lines” that distinguish safe from unsafe exposure levels. Chronic exposure guidance values are set at exposure levels that are expected to be protective over a lifetime of exposure, even for potentially sensitive or susceptible populations. For short-lived compounds such as these phthalates, an exceedance of the corresponding BE value in a single urine sample may or may not reflect continuing elevated exposure. As demonstrated in the limited available datasets, exposures may vary substantially from day to day in an individual (Preau et al., 2007). Thus, occasional exceedances of the BE value in individuals in cross-sectional studies do not imply that adverse health effects are likely to occur, but can serve as an indicator of relative priority for further risk assessment follow-up. A major issue in risk assessment for phthalates is the potential for cumulative exposures and potential adverse effects resulting from exposure to multiple phthalate compounds, even when exposures to individual compounds do not exceed existing exposure guidance values. Howdeshell et al. (2008b) have demonstrated that exposures to multiple endocrine-active phthalate compounds can act additively to produce effects even when the doses of individual compounds are below the respective no-effect levels. As discussed in the introduction to this manuscript, BE values are tied to the underlying exposure guidance values, and to the extent that these values and their underlying risk assessments have not taken into account the potential for cumulative effects across compounds, the BE values also do not address this possibility. One possible approach to using BE values in an assessment of exposures to multiple phthalate compounds is to calculate a cumulative hazard index across compounds, similar to approaches used in risk assessment of contaminated sites under guidance from the USEPA (e.g., USEPA, 2000; Haddad et al., 2000, 2001). This approach sums the ratios of the measured exposures (in this case, biomarker concentrations, CBiomarker) to the tolerable exposure value (in this case, BE values) across compounds:

HI ¼

n X C Biomarker BE i¼1

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A hazard index below 1 suggests that the cumulative exposures of the compounds as fractions of their BE values do not exceed an overall level consistent with the risk assessments underlying the exposure guidance values. Such an approach is likely to be appropriate only when the included compounds affect the same endpoints through similar mechanisms of action, and when doseadditivity (rather than response-additivity, synergism or antagonism) is expected. Thus, with respect to the three compounds included in this manuscript, use of a hazard index approach might be appropriate for DBP and BzBP (and DEHP, included in a companion manuscript), which have been shown to have similar toxicological activity and to act in an additive manner, but inclusion of DEP in such an evaluation would not be appropriate given its lack of activity on these toxicological endpoints. In addition, as discussed above, application to measurements in individuals, or to the tails of distributions in population data sets, is likely to produce misleading results due to the highly transient nature of the phthalate compounds. However, such an approach might be considered in the evaluation of the central tendency of biomonitoring data for a population, and could provide one tool in the assessment of such data. Further discussion of interpretation and communication aspects of the BE values is presented in LaKind et al. (2008) and at www.biomonitoringequivalents.net. 8. 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 the BEs for DEHP. Documentation for the procedure for independent peer-reviews for BEs is available at www.biomonitoringequivalents.net. References Anderson, W.A., Castle, L., Scotter, M.J., Massey, R.C., Springall, C., 2001. A biomarker approach to measuring human dietary exposure to certain phthalate diesters. Food Addit. Contam. 18, 1068–1074. Api, A.M., 2001. Toxicological profile of diethyl phthalate: a vehicle for fragrance and cosmetic ingredients. Food Chem. Toxicol. 39, 97–108. Aylward, L.L., Hays, S.M., Gagné, M., Krishnan, K., 2009. Derivation of Biomonitoring Equivalents for Di(2-ethylhexyl)phthalate (CAS No. 117-81-7). Regul. Toxicol. Pharmacol. 55, 249–258. Barr, D.B., Wilder, L.C., Caudill, S.P., Gonzalez, A.J., Needham, L.L., Pirkle, J.L., 2005. Urinary creatinine concentrations in the U.S. population: implications for urinary biologic monitoring measurements. Environ. Health Perspect. 113, 192– 200. Blount, B.C., Milgram, K.E., Silva, M.J., Malek, N.A., Reidy, J.A., Needham, L.L., Brock, J.W., 2000. Quantitative detection of eight phthalate metabolites in human urine using HPLC-APCI-MS/MS. Anal. Chem. 72, 4127–4134. Calafat, A.M., McKee, R.H., 2006. Integrating biomonitoring exposure data into the risk assessment process: phthalates [diethyl phthalate and di(2-ethylhexyl) phthalate] as a case study. Environ. Health Perspect. 114, 1783–1789. Centers for Disease Control (CDC), 2005. Third National Report on Human Exposure to Environmental Chemicals. NCEH Pub. No. 05-0570. Clewell, R.A., Kremer, J.J., Williams, C.C., Campbell Jr., J.L., Andersen, M.E., Borghoff, S.J., 2008. Tissue exposures to free and glucuronidated monobutylyphthalate in the pregnant and fetal rat following exposure to di-n-butylphthalate: evaluation with a PBPK model. Toxicol. Sci. 103, 241–259. Clewell, R.A., Kremer, J.J., Williams, C.C., Campbell, J.L., Sochaski, M.A., Andersen, M.E., Borghoff, S.J., 2009. Kinetics of selected di-n-butyl phthalate metabolites and fetal testosterone following repeated and single administration in pregnant rats. Toxicology 255, 80–90. EFSA (European Food Safety Authority), 2005a. Opinion of the scientific panel on food additives, flavourings, processing aids and materials in contact with food

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