Regulatory Toxicology and Pharmacology 60 (2011) 165–171

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Regulatory Toxicology and Pharmacology

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Biomonitoring Equivalents for 2,20,4,40,5-pentabromodiphenylether (PBDE-99)

Kannan Krishnan a, Therese Adamou a, Lesa L. Aylward b,⇑, Sean M. Hays c, Christopher R. Kirman d,Andy Nong e

a Université de Montréal, Département de santé environnementale et santé au travail, Montréal, QC, Canadab Summit Toxicology, LLP, Falls Church, VA, USAc Summit Toxicology, LLP, Lyons, CO, USAd Summit Toxicology, LLP, Orange Village, OH, USAe Health Canada, Ottawa, Ontario, Canada

a r t i c l e i n f o

Article history:Available online 3 April 2011

Keywords:Biomonitoring EquivalentsPolybrominated biphenyl ethers2,20 ,4,40 ,5-PentabromodiphenyletherPBDE-99Risk assessmentToxicokinetics

0273-2300/$ - see front matter � 2011 Elsevier Inc. Adoi:10.1016/j.yrtph.2011.03.011

Abbreviations: AF, absorption fraction; BE, BiomBiomonitoring Equivalent point of departure; BMconfidence limit; BW, body weight; CAS, chemical abhalf-life; Kel, elimination rate per day; LOAEL, lowestLOEL, lowest observed effect level; MOE, margin of exadverse effect level; PKC, protein kinase C; PND10, podeparture; RfD, reference dose; T4, thyroxine (thyrodaily intake; TTR, transthyretin; UF, uncertainty faEnvironmental Protection Agency.⇑ Corresponding author. Address: Summit Toxicolo

Falls Church, VA 22044, USA.E-mail address: laylward@summittoxicology.com

a b s t r a c t

Biomonitoring Equivalents (BEs) are defined as the concentration or range of concentrations of a chemicalor its metabolite in a biological medium (blood, urine, or other medium) that is consistent with an exist-ing health-based exposure guideline such as a reference dose (RfD) or tolerable daily intake (TDI). BE val-ues can be used as a screening tool for the evaluation of population-based biomonitoring data in thecontext of existing risk assessments. This study reviews health based risk assessments and exposureguidance values for 2,20 ,4,40 ,5-pentabromodiphenylether (PBDE-99) from Health Canada and the UnitedStates Environmental Protection Agency (US EPA). Toxicokinetic data from laboratory animals andhumans are reviewed. A BE value corresponding to the US EPA RfD is derived here for PBDE-99 basedon the assumption of chronic steady-state exposure, distribution into body lipids, and a previously-esti-mated first-order half-life of elimination of 1040 days. The steady-state lipid-adjusted BERfD is 520 ng/glipid. Sources of uncertainty relating to the underlying toxicokinetic and toxicologic database forPBDE-99 and the simultaneous exposure to multiple PBDE congeners are discussed. The BERfD valuemay be used as a screening tool for evaluation of population biomonitoring data for PBDE-99 in the con-text of the existing US EPA risk assessment and can assist in prioritization of the potential need for addi-tional risk assessment efforts for PBDE-99 relative to other chemicals.

� 2011 Elsevier Inc. All rights reserved.

1. Introduction

The measurement of chemicals and their metabolites in biolog-ical matrices of exposed individuals has to go hand in hand withthe development of tools and reference values for their interpreta-tion. In the absence of such criteria, the human biomonitoring datacannot be put in a health risk context, rather only in terms of expo-sure trends or to confirm the presence or absence of a chemical at agiven point in time. The development of screening criteria could

ll rights reserved.

onitoring Equivalent; BEPOD,DL, benchmark dose lowerstracts services; HL, terminalobserved adverse effect level;posure; NOAEL, no observedstnatal day 10; POD, point ofid hormone); TDI, tolerable

ctor; US EPA, United States

gy, LLP, 6343 Carolyn Drive,

(L.L. Aylward).

most readily be developed with the availability of robust datasetsrelating adverse responses to internal and external doses as wellas the biomarker measures in human populations. The databaseon most chemicals is such that an interim approach, the develop-ment of Biomonitoring Equivalents (BEs), has been proposed, andguidelines for the derivation and communication of these valueshave been developed (Hays et al., 2007, 2008; LaKind et al.,2008). A BE refers to the concentration or range of concentrationsof a chemical or its metabolites in a biological medium that is con-sistent with an existing health-based exposure guidance valuesuch as a reference dose (RfD) or tolerable daily intake (TDI).Chemical-specific pharmacokinetic data are used to estimate bio-marker concentrations that are consistent with the point of depar-ture (POD) used in the derivation of the exposure guidance value(BEPOD), and with the exposure guidance value itself (BE).

The BEs are intended for use as screening tools to provide anassessment of those chemicals that have large, small, or no marginsof safety based on comparison with existing risk assessments andexposure guidance values. BEs are only as robust as the underlyingexposure guidance values and pharmacokinetic data. BE values arenot intended to be diagnostic for potential health effects in

166 K. Krishnan et al. / Regulatory Toxicology and Pharmacology 60 (2011) 165–171

humans, either individually or among a population. BEs have beenderived using available human or animal pharmacokinetic dataand existing exposure guidance values for a number of compoundsincluding acrylamide, cadmium, 2,4-dichlorophenoxyacetic acid,triclosan, bisphenol A and others (Hays et al., 2008; Hays and Ayl-ward, 2009; Krishnan et al., 2010a,b). BEs have not been previouslydeveloped for mixtures or specific congeners of polybrominated di-phenyl ethers (PBDEs).

Polybrominated diphenylethers (molecular formula: C12H10–x,BrxO. x = 1–10; chemical abstracts services [CAS] Registry No.32534-81-9) are a class of chemicals finding use as flame retar-dants in a wide range of products (e.g., circuit boards, television,computers), and occur as 10 possible homologs, with varying num-bers and positions of bromine atoms. Of the polybrominateddiphenylethers, the pentabromodiphenyl ether occurs in as manyas 46 congener forms (Fig. 1) (US EPA, 2008). 2,20,4,40,5-Pentabro-modiphenylether, commonly known as PBDE-99 (molecularweight 564.7; CASRN: 60348-60-9; C12H5Br5O) is frequently re-ported to occur in environmental media and human tissue samples(CDC, 2009; Frederiksen et al., 2009; Meironyté et al., 1999). Eventhough a number of biomonitoring studies around the world havereported levels of PBDEs in serum lipids, very limited effort has fo-cused on developing a reference value for interpreting the availablehuman monitoring data in a public health context. In this regard,NICNAS (2007) proposed a serum lipid concentration of PBDE-99(6600 ng/g lipid) as being associated with a mouse (i.e.,LOAEL = 0.8 mg/kg/d) based on a number of assumptions, i.e., thata constant proportion of an administered dose will partition intoserum, that the blood concentration associated with a unit doseof PBDE-99 will be the same as that of PBDE-47; and that the ser-um lipid content might be equal to 0.5% in neonatal mice.

The present article describes the derivation of BE values forPBDE-99, using chemical-specific information.

2. Available data and approaches

2.1. Available exposure guidance values

An exposure guidance value for PBDE-99 has been published bythe US EPA (2008). Through the use of benchmark dose modelingof the dose–response data from Viberg et al. (2004), US EPA(2008) determined a BMDL of 0.29 mg/kg/d. The POD for thisassessment was based on the neurobehavioural effects seen inadult (8 month old) female mice that had been administered a sin-gle dose of PBDE-99 on PND-10. A total uncertainty factor of 3000was used: 10 for interspecies uncertainty, 10 for intraspecies vari-ability, 10 for database uncertainty (i.e., lack of data on prenataldevelopmental neurotoxicity, a multigeneration reproductivestudy, and conventional subchronic and chronic toxicity studies)and 3 for uncertainty related to extrapolation of effects from a sin-gle dose neurodevelopmental study to a life-time exposure.

Health Canada has not published an exposure guidance valuefor PBDE-99; however it has published a screening level assess-ment in which a LOEL of 0.8 mg/kg/d was identified as the critical

Fig. 1. Generic structure of polybrominated diphenylether compounds. Pentabro-modiphenylether congeners, of which there are as many as 46, have five brominesubstitutions. PBDE-99 has bromine atoms at the 2,20 ,4,40 , and 5 locations.

effect level (Health Canada, 2006). This POD was based on the neu-robehavioral endpoint, consisting of dose- and time-relatedchanges in locomotion, rearing and total activity in mice after asingle oral gavage dose on PND10 and a 5-month post-dosingobservation period (Eriksson et al., 1998, 2001). NICNAS (2007)and European Union (2001) have also developed a point of depar-ture for MOE calculations but have not published a guidance value.

Table 1 summarizes the EPA’s PBDE-99 guidance value, POD, aswell as the applied uncertainty factors.

2.2. Mechanism of action and relevant dose metrics

The effects of PBDE-99 on locomotion and habituation arethought to be related to the impaired development of the choliner-gic system (Viberg et al., 2003, 2004). In this regard, the exposurescenario/window used in the critical toxicity study for PBDE-99(i.e., PNDs 10–14) corresponds to a period of maximum vulnerabil-ity for the developing cholinergic system (US EPA, 2008). Somestudies suggest that PBDE-99 could interact with proteins relatedto the PKC signaling cascade, which is involved in neuronal devel-opment, memory and learning (Kodavanti et al., 2005; US EPA,2008).

Hydroxylated pentaBDE metabolites have been shown in vitroto compete with T4 for binding with high affinity to transthyretin(TTR), a thyroid hormone transport protein (Qiu et al., 2009).Meerts et al. (2000) reported that PBDE-99 can compete with T4-TTR binding only after metabolic conversion induced by rat livermicrosomes. Given the role of thyroid hormones in the regulationof mammalian brain development (Zoeller et al., 2002), PBDE-99and its metabolites might impact the homeostasis of thyroid hor-mones in vivo in rodents or humans but this has not been demon-strated (US EPA, 2008).

Even though experimental evidence supports the interaction ofPBDE-99 with proteins in the neurological system and with theTTR, the relationship of these findings to the reported neurobeha-vioural effects in the critical toxicological study is unclear. Addi-tional data on the relatively weak interactions of PBDE-99 withthe Ah-, estrogen- and androgen- receptor are available but theirimplications for the mode of action of PBDE-99 are unclear (USEPA, 2008). Further, several studies suggest that the toxic effectsof PBDEs could require metabolic activation, but the relative rolesof parent PBDE-99 and its metabolites are not sufficiently eluci-dated (Meerts et al., 2000; Mercado-Feliciano and Bigsby, 2008;Qiu et al., 2009). Therefore, despite the evidence suggesting thatthere are interactions of PBDE-99 and its metabolites with proteinsat the neurological level, there are insufficient data at the presenttime to determine the specific mode of action or identify theappropriate internal dose metrics that are relevant to a specificmode of action for PBDE-99.

2.3. Available pharmacokinetic data

Even though a number of in vitro and in vivo studies relating tothe mechanisms and extent of absorption, distribution, metabo-lism, and excretion of polybrominated diphenylethers are avail-able, there is only limited information on PBDE-99 specifically.The available pharmacokinetic data for PBDE-99 in rats and miceindicate significant interspecies differences (Chen et al., 2006).Whereas the rodent data generally suggest efficient absorption fol-lowing oral dosing and moderate accumulation particularly in adi-pose tissues, the extent of metabolism and excretion would appearto be substantially different between these two species (Chen et al.,2006).

The efficiency of absorption of PBDE-99 by the oral route is esti-mated to be about 85% and 89%, respectively, in rodents and hu-mans (Chen et al., 2006; Darnerud and Risberg, 2006; Geyer

Table 1Health-based exposure reference values and applicable uncertainty factors for PBDE-99.

Organization, criteria (year ofevaluation)

Study description Critical endpoint and dose Uncertainty factors Guidelinevalue

300RfD US EPA (2008) Neurobehavioral mouse study;

Viberg et al. (2004)BMDL of 0.29 mg/kg; decrease inrearing habituation

- 10: intrahuman variability- 3: single-dose exposure to a life-

time exposure,- 10: database deficiencies

0.1 lg/kg/day

RfD, reference dose.

K. Krishnan et al. / Regulatory Toxicology and Pharmacology 60 (2011) 165–171 167

et al., 2004). In vitro microsomal studies suggest a role for CYP1A1/2 and CYP2B2 in the metabolism of PBDE-99 (Chen and Bunce,2003; Chen et al., 2001; Sanders et al., 2005). The primary stepsin the metabolism of PBDEs lead to the formation of arene oxideintermediates, resulting in hydroxylated as well as hydroxylatedand debrominated metabolites. Hydroxylated PBDEs (OH-PBDEs;mainly 50-hydroxy-bromodiphenylether-99 and 60-hydroxy-bro-modiphenylether-99) have also been detected in human serum(Qiu et al., 2009) and recent in vitro (human hepatocyte) studieshave demonstrated the formation of 2,4,5-tribromophenol, mono-hydroxylated metabolites as well as an unidentified tetrabromi-nated metabolite from PBDE-99 (Stapleton et al., 2009). Qiu et al.(2009) suggest mouse-human differences in the pathways of PDBEmetabolism, and reported the occurrence of 50-hydroxy and 60-hy-droxy metabolites in human serum, contrary to bromophenols thatwere found in mice. Interspecies differences in excretion have alsobeen reported, with the rats excreting more radiolabel in the fecesthan in the urine and the opposite pattern observed in mice (Chenet al., 2006). PBDE-99 metabolites are excreted mainly via urine inthe mouse; whereas in the rats the parent chemical is the predom-inant form that is excreted via the feces (Chen et al., 2006).

Consistent with their lipophilicity, the polybrominated diphe-nyl ethers show marked affinity for and accumulation in lipid tis-sues. Animal studies have shown that PBDE congeners includingPBDE-99 are preferentially distributed in adipose tissues (Chenet al., 2006; Darnerud and Risberg, 2006; Hakk et al., 2002; Kuriy-ama et al., 2007). In mice, after a single intravenous dose of 1 mg/kg, the highest concentrations (as reflected by the radiolabel) werefound in fat, skin, muscle and liver (Staskal et al., 2006a, b). Thesetissues accounted for about 40% of the dose administered. On thecontrary, less than 0.5% of the dose was found in brain, lungs, kid-neys and blood. Following an intravenous dose, based on measuresof radioactivity, it was calculated that the cumulative amount ex-creted in urine was 8% of the dose after day 1, increasing to 16%after day 5. Regarding fecal excretion following a single intrave-nous dose of PBDE-99, 21% of the dose was excreted after day 1decreasing to 3.6% on day 5. During this 5-day period, 78–87% ofthe fecal radioactivity was identified as metabolites (calculatedas the sum of metabolites from nonextractable, extractable,water-soluble and lipid-bound fractions) and 10–21% as beingassociated with the parental form of PBDE-99 (Staskal et al.,2006a, b).

Table 2Potential biomarkers of exposure to PBDE-99.

Analyte Medium Advantages

Parent compound PBDE-99 Blood Fairly stable; represents reredistribution of PBDE from

PBDE-99 Urine Non-invasive

Primarymetabolite

Hydroxylatedmetabolites of PBDE-99

Blood –

Urine Specific biomarkers of exp

Neither a compartmental pharmacokinetic model nor a physio-logically-based pharmacokinetic model has been developed tosimulate the kinetics of PBDE-99 in mice, rats or humans. However,estimations of half-lives based on limited available human datahave been reported (Geyer et al., 2004). These authors reportedthe whole body half-lives of PBDE and other brominated flameretardants on the basis of their estimated daily intake and totalbody burden under steady-state conditions in non-occupationallyexposed adult humans. For this purpose, the blood or adipose lipidconcentrations of PBDE in non-occupationally exposed adults inSweden as well as the daily intake of PBDE in adult humans basedon a Swedish market basket survey were obtained from the litera-ture. This analysis indicated that the half-life of PBDE-99 in hu-mans varied from 1.8 to 3.95 years (1040 days) (Geyer et al., 2004).

2.4. Potential biomarkers

The possible biomarkers for PBDE-99 are summarized in Table 2.Even though PBDE-99 in parent form and as metabolite(s) can bemeasured in whole blood, serum, plasma or breast milk (normal-ized to lipid content or not), the parent chemical measure is poten-tially useful as a biomarker of exposure, based on analogy to otherlipophilic chemicals (Verner et al., 2008; Aylward et al., 2010). Theuse of lipid-adjusted concentration of PBDE-99 is expected to be atoxicologically relevant internal dose metric, given the state ofknowledge regarding the physicochemical, metabolic and toxico-logical characteristics of this chemical.

3. BE derivation

Derivation of the BE for PBDE-99 could proceed in one of twoways: (1) through extrapolation from estimated or measured inter-nal doses in the experimental animals at the point(s) of departureunderlying the exposure guidance values; or (2) through estima-tion of human biomarker concentrations (lipid-adjusted PBDE-99concentrations) associated with the human-equivalent POD andthe exposure guidance values using knowledge of the human tox-icokinetics. The estimation of relevant internal tissue concentra-tions at the point(s) of departure is difficult due to the lack ofreliable data and the experimental design, which involved

Disadvantages

cent exposure as well asbody storage

Invasive

% Excreted unchanged in urine of humans notknown but expected to be low

Low level in blood

osure; non-invasive No available data on human urinary kinetics

Table 3Derivation of BE values for PBDE-99 based on US EPA RfD.

BE derivation step US EPARfD

POD, mg/kg-d 0.29Uncertainty Factors (single dose to chronic exposure, and

interspecies; see Table 1):30

Human Equivalent POD, mg/kg-d: 0.0097BEPOD, PBDE concentration in lipid, mg/kg lipid: 51.6UF (intraspecies, database; see Table 1): 100BE, mg/kg lipid: 0.52BE, ng/g lipid: 520

Note: The concentrations in various blood fractions (serum, plasma, or whole blood)are likely to be approximately equivalent as long as the measurements are lipidadjusted. Thus, the values in Table 3 can be applied to interpret the lipid-adjustedPBDE-99 concentrations in whole blood or blood fractions such as serum or plasma.

168 K. Krishnan et al. / Regulatory Toxicology and Pharmacology 60 (2011) 165–171

administration of a single dose of PBDE-99 to neonatal mice. Thus,the second approach is selected here, and is illustrated in Fig. 2.

Given the lipophilicity of PBDE-99 (predicted log Kow = 7.66,Brooke et al., 2009), and in the absence of data suggesting any tis-sue-specific binding proteins, it is reasonable to assume that it isdistributed principally in the lipid stores of the body, a conclusionsupported by data from Hakk et al. (2002). Highly lipophilic com-pounds such as PBDE-99, hexachlorobenzene, and DDT distributeessentially in the volume of lipid in the body. Furthermore, anassumption of simple, first-order elimination kinetics appears tobe appropriate, based on analogy to other highly lipophilic chemi-cals (e.g., Verner et al., 2008). Therefore, the steady-state lipid con-centration (Css) (units) of PBDE-99 can be calculated as follows:

Css ¼ D� BW� AFVd� Kel

ð1Þ

where D, chronic exposure dose (e.g., reference dose or tolerabledaily take, mg/kg-d); BW, body weight (kg) (BW = 70 kg; standardassumption); AF, absorption fraction (unitless) (AF = 0.89, Geyeret al. (2004)); Vd, volume of lipid (0.25 � BW, kg lipid, genericassumption, Aylward et al. (2010)); and Kel, elimination rate (d�1).

The Kel can be computed from the half-life of the chemical.Geyer et al. (2004) computed a terminal half-life (HL) for PBDE-99 on the basis of the body burden and dietary intake data ob-tained from a Swedish market basket study. Accordingly, the aver-age total body half-life of PBDE-99 was reported to be 1040 days.The corresponding Kel is calculated as:

Kel ¼ ln 2HL

ð2Þ

As illustrated in Fig. 2, a BE value corresponding to the US EPA’sRfD was derived as follows:

� The identified point of departure (POD) for the risk assessment(a BMDL; see Table 1 for description) was divided by the iden-tified uncertainty factors for extrapolating from effects seen in asingle dose study to a lifetime exposure (3) and for interspeciesextrapolation (10) (US EPA, 2008) to derive a human-equivalentdose at the POD.� The lipid-adjusted concentration (Css) corresponding to the

identified human-equivalent POD dose (BEPOD) was calculatedthrough use of equation 1 above.� Remaining uncertainty factors for intraspecies (10) extrapola-

tion and database uncertainty factors (10) (US EPA, 2008) wereapplied to the BEPOD to derive the BE in units of mg PBDE-99 per

HumanEquiv. POD

Human PK data (half-life) and Body lipid content

External Dose

Relevant Internal

Dose

Animal

HumanHumanEquiv.BEPOD

MonitoredBiomarker

UF

H

UF

D

Target BE in lipid

UF A

UF

SC

Animal POD

Fig. 2. Approach to deriving BE values for the EPA RfD and POD for PBDE-99. (BE,Biomonitoring Equivalent; POD, point of departure; PK, pharmacokinetic; UF,Uncertainty factor: A, animal to human; SC, subchronic to chronic; H, interindivid-ual; D, database).

kg of lipid, and the result was also converted to units of ng/glipid, which are often used in reporting biomonitoring data forlipophilic chemicals.

Due to the use of lipid-adjusted concentration of PBDE-99,which is expected to be a toxicologically relevant internal dosemetric, the toxicokinetic component of the intraspecies uncer-tainty factor could be considered for replacement with 1 (Hayset al., 2008). However, in this case, there are two sources of uncer-tainty in the calculation that indicate that retention of the toxicoki-netic component is appropriate. First, the estimate of human half-life of elimination is based on limited data, and thus the estimate ofthe BEPOD should be considered to be somewhat uncertain. Second,because the toxicological database for PBDE-99 is limited, it is notknown if the toxicologically active species is the parent compoundor metabolites or both. Thus, retention of the toxicokinetic compo-nent of the intraspecies UF is appropriate.

The steps in the BE derivation and the resulting BEPOD (52 mg/kglipid) and BE (520 ng/g lipid) values for the EPA risk assessment arepresented in Table 3.

4. Discussion

The present study derived BE and BEPOD values for PBDE-99based on the US EPA’s risk assessment. The BE corresponds to thelipid-normalized concentration of PBDE-99, such that it may beused for interpreting serum or plasma values obtained in biomon-itoring programs. PBDE-99 concentrations in various blood frac-tions (serum, plasma, or whole blood) will be approximatelyequivalent as long as the measurements are lipid-adjusted. Thus,the values in Table 3 can be applied to compare with lipid-adjustedbiomarker concentrations in whole blood or blood fractions such asserum or plasma.

BE or BEPOD values associated with the Health Canada’s screen-ing level health risk assessment were not derived in this study, dueto the lack of determination of an exposure guidance value byHealth Canada. However, Equation 1 presented in this article canbe applied to derive BE values corresponding to new exposureguideline values as they become available or get updated as a re-sult of new data or risk assessment practice.

NICNAS (2007) proposed a reference serum lipid concentrationof PBDE-99 associated with a POD chosen from the Eriksson et al.(2001) study (i.e., LOAEL = 0.8 mg/kg/d in mouse). This was basedon three assumptions: (i) a constant proportion of an administereddose will partition into serum, (ii) the blood concentration associ-ated with a unit dose of PBDE-99 will be the same as that of PBDE-47 (during 1–5 days after dosing); and (iii) the serum lipid contentis equal to 0.5% in neonatal mice. Because of the lack of availabilityof pharmacokinetic data associated with the mouse LOAEL, the

K. Krishnan et al. / Regulatory Toxicology and Pharmacology 60 (2011) 165–171 169

data for PBDE-47 were used to derive the serum levels for PBDE-99.Accordingly, NICNAS (2007) developed a value of 6600 ng/g lipid asthe reference blood concentration associated with the LOAEL inneonatal mice (0.8 mg/kg/d, Eriksson et al., 2001) during 1–5 daysafter dosing whereas the present study determined 520 ng/g lipidas the BE. It should be noted that the NICNAS derived the lipid con-centration of PBDE only for the animal POD of 0.8 mg/kg/d, and didnot derive such values for the human equivalent POD. Further, noadjustments for interspecies differences was performed unlike thecurrent study which accounted for the distribution volume of lipidand half-lives in humans and calculated lipid concentrations asso-ciated with human-equivalent POD as well as the exposure guid-ance value. Moreover, in this derivation, no assumption of lipidlevels in neonatal mice (assumed to be 0.5% by NICNAS, 2007)was made. However, the BE value derived here is specific only toPDBE-99 and not for all other PBDE congeners. Since this congeneroccurs as mixtures in exposed people, such biomarker data formultiple congeners would be of relevance in computing a biologi-cal hazard index for common toxicological effects.

4.1. Sources of variability and uncertainty

Several factors contribute to potential uncertainties and limita-tions in the BE values derived in this effort or in their use for inter-pretation of biomonitoring data. Fundamentally, the uncertaintiesand limitations associated with the BE are reflective of the expo-sure guidance value used in its derivation. In the case of PBDE-99, the RfD was developed by US EPA in 2008, when no standardreproductive, developmental (including neurdevelopmental), sub-chronic, or chronic toxicity studies existed in rats or mice. Reflect-ing this state of knowledge at the time of RfD development as wellas the uncertainty related to the principal study (Viberg et al.,2004), EPA indicated low confidence in its RfD. Since then, con-cerns regarding the experimental design, reproducibility and dataanalysis of the Viberg et al. (2004) study have appeared in the lit-erature (Williams and DeSesso, 2010). The protocol used by Viberget al. (2004) was unique in that it did not conform to the existingguidelines for neurotoxicity screening batteries or developmentalneurotoxicity studies (US EPA, 1998b). Also, the dosing regimenof this study did not encompass exposure via gestation and/or lac-tation (US EPA, 1998a); but the fact that effects were observed afteronly a single dose (given during a critical window of developmen-tal susceptibility) argues for the sensitivity of this study despite themethodological limitations and concerns. The guideline values willevolve with the science and supporting data. In the case of PBDE,there are a number of newer studies examining commercial PBDEmixtures (e.g., DE 71) rather than the specific congener of interest(e.g., Alonso et al., 2010; Kodavanti et al., 2010; Dunnick and Nys-ka, 2009), and some are of limited power (Mazdai et al., 2003;Julander et al., 2005; Herbstman et al., 2008, 2010; Turyk et al.,2008; Roze et al., 2009; Chevrier et al., 2010).

Another factor that contributes to uncertainties and limitationsof BE derivation is the underlying pharmacokinetic data. There wasno pharmacokinetic data, model or clearance parameter collectedas part of the critical toxicological study used to derive RfD (Viberget al., 2004). However, there are data on the total radioactivityassociated with PBDE-99 in neonatal mouse brain for similar pro-tocols (Eriksson et al., 2002). The BEs could then have been derivedfrom the internal dose in exposed animals. In this regard, Erikssonet al. (2002) reported that the radioactivity found in brain of neo-natal mice was about 0.37–0.51% of the dose administered at 24 hpost-dosing, and it declined to 0.13–0.28% of the administereddose at 7 days after dosing. The calculated total concentration ofPBDE-99 in the brain of mice was 7.2–10.7 pmol/g on postnatalday 10, at 24 h after a single oral dose of 0.8 mg/kg, based on re-tained radioactivity (Eriksson et al., 2002). Considering the brain

as the target organ, this option could have represented a relevantchoice. However, neither the brain content of lipids on PND10,nor the relative proportions of metabolites vs parent chemical (freevs total) was known. Adding further to these sources of uncertaintyare uncertainties regarding the body lipid content in mice onPND10 as well as the relative distribution of PBDE-99 in brain com-pared to the rest of the body including adipose tissues. Since thefirst three to four postnatal weeks correspond to the most inten-sive period for brain lipid synthesis in mice (Eriksson et al.,2002), approximations of these critical parameters would onlyintroduce significant uncertainty in the BE derivation for PBDE-99. The ambiguity and uncertainties associated with such calcula-tions of PBDE-99 dose to brain in the mouse and its extrapolationto humans has previously been debated (Eriksson et al., 2002; Eri-ksson and Viberg, 2004; Vijverberg and van den Berg, 2004; Wil-liams and DeSesso, 2010).

In the present derivation, the relationship between the bio-marker and external dose in humans was based on steady-state ki-netic considerations. Accordingly, an elimination rate constant andvolume of distribution were used to compute BEs. The eliminationconstant (Kel) for PBDE-99 in humans used in the computation ofthe BE corresponds to the mean value obtained from the range re-ported by Geyer et al. (2004). These authors determined the half-life of PBDE-99 in humans, using two approaches: (i) on the basisof human blood concentration and dietary uptake data in Swedishpopulation, and (ii) scaling of the half-life in the rat on the basis ofan empirical extrapolation approach. The human data-based ap-proach yielded half-lives ranging from 1.8 to 4 years, with a meanvalue of 2.9 years. The human half-life estimated from rat data (3.5to 7.2 years) was calculated using the following empirical relation-ship, which was based on a number of persistent organic chemicals(Geyer et al., 2004):

Log human half-life ¼ ð1:34� Log rat half-lifeÞ þ 1:25 ð3Þ

The present study used the Kel derived from the human dataand not the value extrapolated from the rat data. Use of the longerhalf-life estimates based on the rat data would result in higher BEvalues than those derived here. Given the reported variability inthe Kel values derived from human data and a Swedish market bas-ket study (which may not be reflective of other populations), aswell as interindividual variability in lipid content (e.g., Brownet al., 1997), the use of a full intraspecies uncertainty factor of 10reflecting both toxicodynamic and toxicokinetic uncertainties –as in the US EPA risk assessment – was applied during BE deriva-tion. Other potential sources of variability include possible ageand gender differences; however, the experimental studies donot suggest any gender-related differences in PBDE-99 dispositionor neurobehavioral responses (US EPA, 2008).

4.2. Confidence assessment

Two main elements in the assessment of confidence in the de-rived BE values (Hays et al., 2008) are: (i) the robustness of theavailable pharmacokinetic data and models, and (ii) the character-ization or understanding of the relationship between the measuredbiomarker and the critical or relevant target tissue dose metric. Asdiscussed above, there are limitations regarding the pharmacoki-netic database in the mouse and uncertainties in the humanparameters used in the BE derivation for PBDE-99. The confidencein the relevance of PBDE-99 in lipid as a marker for brain concen-tration of the relevant dose metric (i.e., free or total concentrationof the parent form or a relevant metabolite) and therefore for neu-robehavioral toxicity is low-medium. Thus, the assessment of theconfidence level in the derived BE values based on these two fac-tors is as follows:

170 K. Krishnan et al. / Regulatory Toxicology and Pharmacology 60 (2011) 165–171

� Robustness of pharmacokinetic data: LOW-MEDIUM.� Relevance of biomarker to relevant dose metrics: LOW-

MEDIUM.

4.3. Interpretation of biomonitoring data using BE values

Biomonitoring data for PBDE-99 reported on a lipid-adjustedbasis from measurements in serum or plasma can be comparedto the BE values developed in this study. Comparison of lipid-nor-malized concentrations as reported in biomonitoring studies to theBE values can provide an initial evaluation of whether measuredlevels in a given study are of low, medium, or high priority for riskassessment follow-up. Measured biomarker values in excess of thehuman equivalent BEPOD indicate a high priority for risk assess-ment follow-up. Values within the range of the estimated BE val-ues suggest a medium priority for risk assessment follow-up,while those below the BE values suggest lower priority (LaKindet al., 2008). Based on the results of such comparisons betweenbiomonitoring data for a population and the BE values, an evalua-tion can be made of the need for additional studies on exposurepathways, potential adverse health effects, other aspects affectingexposure or risk, or other risk management activities for eitherthe population as a whole or for subsections of the population thatmay exhibit biomarker concentrations in the range of higher prior-ity for follow-up. In this regard, the BE values do not representdiagnostic criteria and cannot be used to evaluate the likelihoodof an adverse health effect in an individual or even among a popu-lation. Further discussion of interpretation and communicationsaspects of BE values is presented in LaKind et al. (2008).

5. Conflict of interest statement

The authors declare they have no conflicts of interest.

Acknowledgments

Funding for this project was provided under a contract fromHealth Canada. This BE derivation has undergone an independentpeer-review to assure the methods employed here are consistentwith the guidelines for derivation (Hays et al., 2008) and commu-nication (LaKind et al., 2008) of Biomonitoring Equivalents and thatthe best available chemical-specific data was used in calculatingthe BEs. We thank the various reviewers for their insightfulsuggestions.

References

Alonso, V., Linares, V., Belles, M., Albina, M.L., Pujol, A., Domigo, J.L., 2010. Effects ofBDE-99 on hormone homeostasis and biochemical parameters in adult malerats. Food Chem. Toxicol. 48 (8–9), 2206–2211.

Aylward, L.L., Hays, S.M., Gagné, M., Nong, A., Krishnan, K., 2010. Biomonitoringequivalents for hexachlorobenzene. Regul. Toxicol. Pharmacol. 58, 25–32.

Brooke, D.N., Burns, J., Crookes, M.J., Dungey, S.M., 2009. Using science to create abettezr place. Environmental risk evaluation report. Decabromodiphenyl ether(CAS No. 1163-19-5). Environment Agency, Bristol, UK.

Brown, R.P., Delp, M.D., Lindstedt, S.L., Rhomberg, L.R., Beliles, R.P., 1997.Physiological parameter values for physiologically based pharmacokineticmodels. Toxicol. Ind. Health 13, 407–484.

Centers for Disease Control and Prevention (CDC), 2009. Fourth National Report onHuman Exposure to Environmental Chemicals. Department of Health andHuman Service. Atlanta, GC. Available from: <http://www.cdc.gov/exposurereport> (accessed 29.11.2010).

Chen, G., Bunce, N.J., 2003. Polybrominated diphenyl ethers as Ah receptor agonistsand antagonists. Toxicol. Sci. 76, 310–320.

Chen, G., Konstantinov, A.D., Chittim, B.G., Joyce, E.M., Bols, N.C., Bunce, N.J., 2001.Synthesis of polybrominated diphenyl ethers and their capacity to induceCYP1A by the Ah receptor mediated pathway. Environ. Sci. Technol. 35, 3749–3756.

Chen, L.J., Lebetkin, E.H., Sanders, J.M., Burka, L.T., 2006. Metabolism and dispositionof 2,20 ,4,40 ,5-pentabromodiphenyl ether (BDE99) following single or repeatedadministration to rats and mice. Xenobiotica 36, 515–534.

Chevrier, J., Harley, K.G., Bradman, A., Gharbi, M., Sjodin, A., Eskanzi, B., 2010.Polybrominated diphenylether (PBDE) flame retardants and thyroid hormoneduring pregnancy. Environ. Health Perspect. 118, 1444–1449.

Darnerud, P.O., Risberg, S., 2006. Tissue localisation of tetra- andpentabromodiphenyl ether congeners (BDE-47, -85 and -99) in perinatal andadult C57BL mice. Chemosphere 62, 485–493.

Dunnick, J.K., Nyska, A., 2009. Characterization of liver toxicity in F344/N rats andB6C3F1 mice after exposure to a flame retardant containing lower molecularweight polybrominated diphenyl ethers. Exp. Toxicol. Pathol. 61 (1), 1–12.

Eriksson, P., Jakobsson, E., Fredriksson, A., 1998. Developmental neurotoxicity ofbrominated flame-retardants, polybrominated diphenyl ethers and tetrabromo-bis-phenol A. Organohalogen Compd. 35, 375–377.

Eriksson, P., Jakobsson, E., Fredriksson, A., 2001. Brominated flame retardants: anovel class of developmental neurotoxicants in our environment? Environ.Health Perspect. 109, 903–908.

Eriksson, P., Viberg, H., 2004. Letter to the editor. Toxicol. Sci. 79, 207–208.Eriksson, P., Viberg, H., Jakobsson, E., Orn, U., Fredriksson, A., 2002. A brominated

flame retardant, 2,20 ,4,40 ,5-pentabromodiphenyl ether: uptake, retention, andinduction of neurobehavioral alterations in mice during a critical phase ofneonatal brain development. Toxicol. Sci. 67, 98–103.

Frederiksen, M., Vorkamp, K., Thomsen, M., Knudsen, L.E., 2009. Human internal andexternal exposure to PBDEs – a review of levels and sources. Int. J. Hyg. Environ.Health 212, 109–134.

Geyer, H.J., Schramm, K.W., Darnerud, P.O., Aune, M., Feicht, E.A., Fried, K.W., 2004.Terminal elimination half-lives of the brominated flame retardants TBBPA, HBCD,and lower brominated PBDEs in humans. Organohalogen Compd. 66, 3867–3872.

Hakk, H., Larsen, G., Klasson-Wehler, E., 2002. Tissue disposition, excretion andmetabolism of 2,20 ,4,40 ,5-pentabromodiphenyl ether (BDE-99) in the maleSprague-Dawley rat. Xenobiotica 32, 369–382.

Hays, S.M., Aylward, L.L., 2009. Using biomonitoring equivalents to interpret humanbiomonitoring data in a public health risk context. J. Appl. Toxicol. 29, 275–288.

Hays, S.M., Aylward, L.L., LaKind, J.S., Bartels, M.J., Barton, H.A., Boogaard, P.J., Brunk,C., DiZio, S., Dourson, M., Goldstein, D.A., Lipscomb, J., Kilpatrick, M.E., Krewski,D., Krishnan, K., Nordberg, M., Okino, M., Tan, Y.M., Viau, C., Yager, J.W., 2008.Guidelines for the derivation of biomonitoring equivalents: report from thebiomonitoring equivalents expert workshop. Regul. Toxicol. Pharmacol. 51, S4–S15.

Hays, S.M., Becker, R.A., Leung, H.W., Aylward, L.L., Pyatt, D.W., 2007. Biomonitoringequivalents: a screening approach for interpreting biomonitoring results from apublic health risk perspective. Regul. Toxicol. Pharmacol. 47, 96–109.

Health Canada, 2006. State of the science report for a screening health assessmentpolybrominated diphenyl ethers (PBDEs) [Tetra-, Penta-, Hexa-, Hepta-, Octa-,Nona- and Deca-Congeners] [numéros CAS 40088-47-9, 32534-81-9, 36483-60-0, 68928-80-3, 32536-52-0, 63936-56-1, 1163-19-5], p. 35.

Herbstman, J.B., Sjödin, A., Apelberg, B.J., Witter, F.R., Halden, R.U., Patterson, D.G.,Panny, S.R., Needham, L.L., Goldman, L.R., 2008. Birth delivery mode modifiesthe associations between prenatal polychlorinated biphenyl (PCB) andpolybrominated diphenyl ether (PBDE) and neonatal thyroid hormone levels.Environ. Health Perspect. 116, 1376–1382.

Herbstman, J.B., Sjodin, A., Kurzon, M., Lederman, S.A., Jones, R.S., Rauh, V., Needham,L.L., Tang, D., Niedzwiecki, M., Wang, R.Y., Perera, F., 2010. Prenatal exposure toPBDEs and neurodevelopment. Environ. Health Perspect. 118 (5), 712–719.

Julander, A., Karlsson, M., Hagstrom, K., Ohlson, C.G., Engwall, M., Bryngelsson, I.L.,Westberg, H., van Bavel, B., 2005. Polybrominated diphenyl ethers-plasmalevels and thyroid status of workers at an electronic recycling facility. Int. Arch.Occup. Environ. Health 78, 584–592.

Kodavanti, P.R., Ward, T.R., Ludewig, G., Robertson, L.W., Birnbaum, L.S., 2005.Polybrominated Diphenyl Ether (PBDE) effects in rat neuronal cultures: 14C-PBDE accumulation, biological effects, and structure–activity relationships.Toxicol. Sci. 88, 181–192.

Kodavanti, P.R., Coburn, C.G., Moser, V.C., MacPhail, R.C., Fenton, S.E., Stoker, T.E.,Rayner, J.L., Kannan, K., Birnbaum, L.S., 2010. Developmental exposure to acommercial PBDE mixture, DE-71: neurobehavioral, hormonal, andreproductive effects. Toxicol. Sci. 116 (1), 297–312.

Krishnan, K., Gagné, M., Nong, A., Aylward, L.L., Hays, S.M., 2010a. Biomonitoringequivalents for triclosan. Regul. Toxicol. Pharmacol. 58, 10–17.

Krishnan, K., Gagné, M., Nong, A., Aylward, L.L., Hays, S.M., 2010b. Biomonitoringequivalents for bisphenol A (BPA). Regul. Toxicol. Pharmacol. 58, 18–24.

Kuriyama, S.N., Wanner, A., Fidalgo-Neto, A.A., Talsness, C.E., Koerner, W., Chahoud,I., 2007. Developmental exposure to low-dose PBDE-99: tissue distribution andthyroid hormone levels. Toxicology 242, 80–90.

LaKind, J.S., Aylward, L.L., Brunk, C., DiZio, S., Dourson, M., Goldstein, D.A., Kilpatrick, M.E.,Krewski, D., Bartels, M.J., Barton, H.A., Boogaard, P.J., Lipscomb, J., Krishnan, K.,Nordberg, M., Okino, M., Tan, Y.M., Viau, C., Yager, J.W., Hays, S.M., 2008. Guidelinesfor the communication of biomonitoring equivalents: report from the biomonitoringequivalents expert workshop. Regul. Toxicol. Pharmacol. 51, S16–S26.

Mazdai, A., Dodder, N.G., Abernathy, M.P., Hites, R.A., Bigsby, R.M., 2003.Polybrominated diphenyl ethers in maternal and fetal blood samples. Environ.Health Perspect. 111 (9), 1249–1252.

Meerts, I.A., van Zanden, J.J., Luijks, E.A., van Leeuwen-Bol, I., Marsh, G., Jakobsson,E., Bergman, A., Brouwer, A., 2000. Potent competitive interactions of somebrominated flame retardants and related compounds with human transthyretinin vitro. Toxicol. Sci. 56, 95–104.

Meironyté, D., Norén, K., Bergman, A., 1999. Analysis of polybrominated diphenylethers in Swedish human milk. A time-related trend study, 1972–1997. J.Toxicol. Environ. Health A 58, 329–341.

K. Krishnan et al. / Regulatory Toxicology and Pharmacology 60 (2011) 165–171 171

Mercado-Feliciano, M., Bigsby, R.M., 2008. Hydroxylated metabolites of thepolybrominated diphenyl ether mixture DE-71 are weak estrogen receptor-alpha ligands. Environ. Health Perspect. 116, 1315–1321.

National Industrial Chemicals Notification and Assessment Scheme (NICNAS), 2007.Interim public health risk assessment of certain PBDE congeners. AustralianGovernment Department of Health and Aging. Available from: <http://www.nicnas.gov.au/publications/car/other/final%20interim%20report%20-%20march.pdf> (accessed 29.11.2010).

Roze, E., Meijer, L., Bakker, A., Van Braeckel, K.N., Sauer, P.J., Bos, A.F., 2009. Prenatalexposure to organohalogens, including brominated flame retardants, influencesmotor, cognitive, and behavioral performance at school age. Environ. HealthPerspect. 117, 1953–1958.

Qiu, X., Bigsby, R.M., Hites, R.A., 2009. Hydroxylated metabolites of polybrominateddiphenyl ethers in human blood samples from the United States. Environ.Health Perspect. 117, 93–98.

Sanders, J.M., Burka, L.T., Smith, C.S., Black, W., James, R., Cunningham, M.L., 2005.Differential expression of CYP1A, 2B, and 3A genes in the F344 rat followingexposure to a polybrominated diphenyl ether mixture or individualcomponents. Toxicol. Sci. 88, 127–133.

Stapleton, H.M., Kelly, S.M., Pei, R., Letcher, R.J., Gunsch, C., 2009. Metabolism ofpolybrominated diphenyl ethers (PBDEs) by human hepatocytes in vitro.Environ. Health Perspect. 117, 197–202.

Staskal, D.F., Diliberto, J.J., Birnbaum, L.S., 2006a. Disposition of BDE 47 indeveloping mice. Toxicol. Sci. 90, 309–316.

Staskal, D.F., Hakk, H., Bauer, D., Diliberto, J.J., Birnbaum, L.S., 2006b. Toxicokineticsof polybrominated diphenyl ether congeners 47, 99, 100, and 153 in mice.Toxicol. Sci. 94, 28–37.

Turyk, M.E., Persky, V.W., Imm, P., Knobeloch, L., Chatterton, R., Anderson, H.A.,2008. Hormone disruption by PBDEs in adult male sport fish consumers.Environ. Health Perspect. 116, 1635–1641.

United States Environmental Protection Agency (US EPA), 1998a. Guidelines forNeurotoxicity Risk Assessment. Federal Register 63, 26926–26954. Availableonline at: <http://www.epa.gov/ncea/raf/rafguid.htm>.

United States Environmental Protection Agency (US EPA), 1998b. Health effects testguidelines: neurotoxicity screening battery. Office of Prevention, Pesticides andToxic Substances, Washington, DC; OPPTS 870.6200; EPA 712-C-98-238.

United States Environmental Protection Agency (US EPA), 2008. Toxicologicalreview of 2,20 ,4,40 ,5-Pentabromodiphenyl Ether (BDE-99) (CAS No. 60348-60-9).In support of summary information on the integrated risk information system(IRIS). Washington, DC.

Verner, M.A., Charbonneau, M., Lopez-Carrillo, L., Haddad, S., 2008. Physiologicallybased pharmacokinetic modeling of persistent organic pollutants for lifetimeexposure assessment: a new tool in breast cancer epidemiologic studies.Environ. Health Perspect. 116, 886–892.

Viberg, H., Fredriksson, A., Eriksson, P., 2004. Investigations of strain and/or genderdifferences in developmental neurotoxic effects of polybrominated diphenylethers in mice. Toxicol. Sci. 81, 344–353.

Viberg, H., Fredriksson, A., Jakobsson, E., Orn, U., Eriksson, P., 2003. Neurobehavioralderangements in adult mice receiving decabrominated diphenyl ether (PBDE209) during a defined period of neonatal brain development. Toxicol. Sci. 76,112–120.

Vijverberg, H.P.M., van den Berg, M., 2004. Letter to the editor. Toxicol. Sci. 79, 205–206.

Williams, A.L., DeSesso, J.M., 2010. The potential of selected brominated flameretardants to affect neurological development. J. Toxicol. Environ. Health B Crit.Rev. 13 (5), 411–448.

Zoeller, T.R., Dowling, A.L., Herzig, C.T., Iannacone, E.A., Gauger, K.J., Bansal, R., 2002.Thyroid hormone, brain development and the environment. Environ. HealthPerspect. 110, 355–361.