8 - geens et al. (2012)

Food and Chemical Toxicology 50 (2012) 3725–3740

Contents lists available at SciVerse ScienceDirect

Food and Chemical Toxicology

journal homepage: www.elsevier .com/locate / foodchemtox

Review

A review of dietary and non-dietary exposure to bisphenol-A

Tinne Geens a,k, Dominique Aerts b,k, Carl Berthot c,k, Jean-Pierre Bourguignon d,k, Leo Goeyens e,k,Philippe Lecomte f,k, Guy Maghuin-Rogister g,k, Anne-Madeleine Pironnet h,k, Luc Pussemier i,k,Marie-Louise Scippo g,k, Joris Van Loco j,k, Adrian Covaci a,k,⇑a Toxicological Centre, University of Antwerp, Universiteitsplein 1, Antwerp, Belgiumb Federal Public Service of Health, Food Chain Safety and Environment, Place Victor Horta 40/10, 1060 Brussels, Belgiumc DG Animals, Plants and Food, FPS Health, Food Chain Safety and Environment, Eurostation, Place Victor Horta 40/10, 1060 Brussels, Belgiumd Department of Pediatrics, University of Liège, CHU ND des Bruyères, B4030 Chénée, Belgiume Analytical and Environmental Chemistry, University of Brussels, Pleinlaan 2, 1050 Brussels, Belgiumf Center for Education and Research on Macromolecules (CERM), University of Liege, B6, Sart-Tilman, Belgiumg Department of Food Sciences, University of Liege, BatB43b, Sart Tilman, Belgiumh Superior Health Council, Rue de l’Autonomie 4, 1070 Brussels, Belgiumi Veterinary and Agrochemical Research Center (CODA-CERVA), Leuvensesteenweg 17, 3080 Tervuren, Belgiumj Scientific Institute of Public Health, Department of Food, Medicines and Consumer Safety, Rue Juliette Wytsmanstraat 14, 1050 Brussels, Belgiumk Belgian Superior Health Council, FPS Health, Food Chain Safety and Environment, Rue de l’Autonomie 4, 1070 Brussels, Belgium

a r t i c l e i n f o

Article history:Received 1 April 2012Accepted 28 July 2012Available online 4 August 2012

Keywords:Bisphenol-AReviewHuman exposureFood sourcesNon-food sourcesAlternatives

0278-6915/$ - see front matter � 2012 Elsevier Ltd. Ahttp://dx.doi.org/10.1016/j.fct.2012.07.059

⇑ Corresponding author. Address: Toxicological CeUniversiteitsplein 1, 2610 Wilrijk, Belgium. Tel.: +322722.

E-mail address: [email protected] (A. Covaci)

a b s t r a c t

Due to the large number of applications of bisphenol-A (BPA), the human exposure routes are multiple.We aimed to review shortly the food and non-food sources of BPA, and to evaluate their contribution tothe human exposure. Food sources discussed here include epoxy resins, polycarbonate and other appli-cations, such as paperboard and polyvinylchloride materials. Among the non-food sources, exposuresthrough dust, thermal paper, dental materials, and medical devices were summarized. Based on the avail-able data for these exposure sources, it was concluded that the exposure to BPA from non-food sources isgenerally lower than that from exposure from food by at least one order of magnitude for most studiedsubgroups. The use of urinary concentrations from biomonitoring studies was evaluated and the back-calculation of BPA intake seems reliable for the overall exposure assessment. In general, the total expo-sure to BPA is several orders of magnitude lower than the current tolerable daily intake of 50 lg/kg bw/day. Finally, the paper concludes with some critical remarks and recommendations on future humanexposure studies to BPA.

� 2012 Elsevier Ltd. All rights reserved.

Contents

1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3726
1.1. Properties and applications of bisphenol-A . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37261.2. Toxicity of bisphenol-A . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37271.3. European legislation regarding migration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37271.4. Aims of the review . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3727
2. Food exposure to bisphenol-A . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3727
2.1. Epoxy resins. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3728
2.1.1. Migration. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37282.1.2. Levels. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3728

2.2. Polycarbonate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3729
2.2.1. Migration and hydrolysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37292.2.2. Levels. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3730
2.3. Other food contact applications . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3730

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.

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Fig. 1.and E.

3726 T. Geens et al. / Food and Chemical Toxicology 50 (2012) 3725–3740

2.4. Intake estimation from food exposure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3730
3. Non-food sources to bisphenol-A . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3731
3.1. Dust . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37313.2. Thermal paper . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37313.3. Other types of papers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37323.4. Dental materials . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37323.5. Medical devices and healthcare applications. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37323.6. Other non-food sources . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3733

4. Toxicokinetics and metabolism of bisphenol-A . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37335. Human biomonitoring . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3733

5.1. Urinary BPA (ng/mL) � urinary output (mL/day)/body weight (kg) = ng BPA/kg/day . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3735
6. Overall estimation of exposure to bisphenol-A . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37357. Epidemiological studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37368. General discussion and recommendations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3736
Conflict of Interest . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3737Acknowledgments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3737References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3737

1. Introduction

1.1. Properties and applications of bisphenol-A

Bisphenol-A (BPA) [4,40-dihydroxy-2,2-diphenylpropane, CAS80-05-7] (Fig. 1A) is an industrial chemical synthesized by conden-sation of two phenol groups and one acetone molecule. While BPAwas first synthesized in 1891, its estrogenic properties have beenhypothesized in the 1930s (Dodds and Lawson, 1938). Since 1940,BPA was predominantly used as (1) a monomer in the manufactur-ing of polymers such as polycarbonate, PC (Fig. 1B), epoxy resins(Fig. 1C), polysulfone, or polyacrylate, (2) as an antioxidant andinhibitor of end of polymerization in polyvinyl chloride plastics(PVC) and (3) as a precursor for the synthesis of the flame retardanttetrabromobisphenol-A (Geens et al., 2011). Polycarbonate is

A. Chemical structure of bisphenol-A; B. Synthesis of polycarbonate from bisphenChemical structure of bisphenol-S.

currently used in materials intended to come into contact withfood, e.g., reusable plastic bottles, feeding-bottles, plates, goblets,cups, microwave ovenware, storage containers, etc., whereas theepoxy resins are used for internal coating of food and beverage cans(EFSA, 2006). However, only 3% of the produced polycarbonate, aswell as 10% of the epoxy resins, is used in materials intended tocome into contact with foodstuffs (Plastics Europe, 2007). Thereare several other uses of polycarbonates, epoxy resins, polysulfone,and polyacrylates such as sunglasses, building materials, CD-ROM,medical devices, dental materials, etc. BPA is also used in thermalpaper (Geens et al., 2011). For a review of all applications of poly-carbonate and epoxy resins, see ANSES (2011a).

Besides BPA, many bisphenol analogues can be obtained by con-densation of a ketone or an aldehyde with phenols with either var-iation in the carbonyl derivative or in the substituents on the

ol-A; C. Chemical structure of an epoxy resin; D. Chemical structure of bisphenol-F;

T. Geens et al. / Food and Chemical Toxicology 50 (2012) 3725–3740 3727

aromatic ring. Although a large number of compounds can be ob-tained by this route, many are too expensive for an industrialapplication. The toxicities of most of these compounds are notknown, especially when synthesized in research laboratories. Forinstance, a systematic research in SciFinder allows finding 28746compounds inserting the OH–Ar–CH2–Ar–OH subunit. Amongthem, (only) 1010 are commercially available. From these bisphe-nols, bisphenol-F (BPF) (bis(4-hydroxyphenyl)-methane) (Fig. 1D)is increasingly used, because of its lower viscosity and better resis-tance against solvents than the BPA epoxy resin (Danzl et al., 2009).Bisphenol-S (BPS) (4,4-dihydroxy-phenylsulfone) (Fig. 1E) can alsobe used as a monomer in the plastic industry.

1.2. Toxicity of bisphenol-A

Since BPA showed estrogenic properties in a large number ofstudies (reviewed by Chapin et al. (2008)), it is described as anendocrine disruptor chemical (EDC). It is in particular able to bindand activate the human estrogen receptor (the estrogenic proper-ties of BPA were already shown in 1938 by Dodds and Lawson),but with a capacity 1000–5000 times less than the endogenous17-b-oestradiol (FASFC, 2009; Roy et al., 2009). BPF and BPS alsodisplay estrogenic properties (Chen et al., 2002). Moreover, BPAhas been shown to interact with other endocrine receptors, e.g.,thyroid hormone receptors, peroxysome proliferator-activatedreceptor gamma (Diamanti-Kandarakis et al., 2009). BPA was clas-sified as a reproductive toxic substance of category 3 as an alarm-ing substance for the human fertility (INSERM, 2010).

The EFSA published a first risk assessment on BPA in 2006,based on a tolerable daily intake (TDI) of 50 lg/kg body weight/day, and concluded that human exposure through food is lowerthan the TDI, even for babies and young children (EFSA, 2006). Inthe light of new published data, the EFSA has concluded in 2008and 2010 that there was no need to decrease the TDI (EFSA,2008, 2010). However, until now, only the exposure to BPAthrough food has been documented. Yet, as indicated above, BPAcan be found in a large number of non-food applications, whichnecessitates a newer look at the BPA exposure routes posing healthrisks.

Several scientists, including the experts from the French Agencyfor Food, Environmental, and Occupational Health and Safety(ANSES), did not agree with the use of TDI for risk assessmentson EDCs (ANSES, 2010; vom Saal and Hughes, 2005). Their opinionis based on the effects of EDCs observed at low doses, non-mono-tonic dose–response curves, as well as on effects occurring fromvery specific windows of exposure (in particular, early in uteroexposure) (Diamanti-Kandarakis et al., 2009).

The toxicity of BPA has once more been reviewed in a recentANSES report (ANSES, 2011b), with a special focus on effects ofBPA at low dose, e.g. a dose below the NOAEL of 5 mg/kg bodyweight/day from which the current TDI-value of 50 lg/kg bodyweight/day has been derived by EFSA (2006). The French expertshave reviewed the state-of-the-art regarding effects of BPA onthe male and female reproductive system, on brain and behavior,on metabolism and cardiovascular system, on thyroid, on the im-mune system, on intestine, prostate and breast (ANSES, 2011b).In general, it is not possible to conclude definitively on the effectson humans, because of heterogeneous and sometimes poor epide-miological data. ‘‘Suspected’’ negative effects in humans are de-scribed on the maturation of oocytes, the cardiovascular systemand the development of diabetes. The feasibility of human epide-miological studies, however, remains questionable for several rea-sons. BPA cannot be isolated from the mixture of EDCs to whichhumans are exposed to. There is virtually no ‘‘control’’ or unex-posed population due to ubiquity of BPA. There could be an intervalof several decades between the fetal and early postnatal critical

windows of exposure and delayed BPA effects, such as metabolicsyndrome in adulthood. Finally, due to the short half-life of BPA,the urinary levels provide only an estimate of exposure duringthe few previous days (Dekant and Völkel, 2008). In animals, pre-natal or post-natal exposures to low doses of BPA have an effecton different physiological systems. These systems are the maleand female reproductive systems (increase of ovarian cysts, hyper-plasia of the endometrium, precocious puberty, and in adults, de-crease of the sperm production), the brain (neurogenesis andsynaptogenesis), the lipid metabolism and sensitivity to insulin,the immune system, the breast development (hyperplasia) (ANSES,2011b).

1.3. European legislation regarding migration

Since chemical substances can be released from plastic materi-als and articles intended to come into contact with food (Barnes,2006), migration limits are mentioned in the European Legislationfor all permitted substances in plastic materials. For BPA, the spe-cific migration limit (SML) is fixed to 0.6 mg/kg food since 2004and has not been changed, except for baby bottles, for which BPAis banned in EU since 2011. BPS has a SML of 0.05 mg/kg food(EC, 2011a), while BPF is not allowed in plastic materials intendedto come in contact with food by European law. For plastic materialsin contact with food, SMLs have been fixed assuming that 1 kg offood is consumed daily by a person of 60 kg for a lifetime exposure.

For the control of the migration of these chemical substancesfrom the material to the food, it is necessary to distinguish be-tween materials and articles that are already in contact with foodand those which are not yet. For both groups, guidelines are givenin the Regulation (EU) N�10/2011 (EC, 2011a).

Briefly, for materials in contact with food, the migration is mea-sured in food. The contact between the material and the food has tobe ended before the expiration date. The foodstuff has to be pre-pared in accordance with cooking instructions on the package.The parts of food not intended for human consumption are then re-moved and discarded, and the remainder food is homogenized andanalyzed for the presence of the compound of interest, to check thecompliance with the SML.

For materials and articles not in contact with food, a series oftest media are used, simulating the transfer of substances fromthe packaging material to food. These media should represent themain physicochemical properties of food. When using these simu-lants, the standardized time and temperature of the assay must, asfar as possible, reflect the potential migration of the target sub-stance in the food (Grob et al., 2006). These simulants are then ana-lyzed for the presence of the compound of interest, to check thecompliance with the SML.

1.4. Aims of the review

The aims of the present review were to summarize the recentliterature (mostly after 2009 and until December 2011) regardingthe food and non-food sources of BPA (with emphasis on the lat-ter). The compiled information was further used to evaluate thecontribution of various exposure sources to the total human expo-sure. Finally, the authors tried to identify the gaps and needs thatare required for a valuable risk assessment of BPA.

2. Food exposure to bisphenol-A

Generally, food, and especially canned food, is considered as thepredominant source of BPA. Contamination of food with BPA isusually caused by contact with food packaging materials contain-ing epoxy resins and PC. Epoxy resins, as well as PVC organosols

Table 1Overview of BPA in canned food samples and canned beverages.

Country Sample size Detection freq. (%) Range Refs.

Canned food (ng/g)US 78 91 <2–730 Noonan et al. (2011)US 97 59 <0.2–65 Schecter et al. (2010)Canada 78 99 <0.6–534 Cao et al. (2010)Japan 48 92 <1–842 Sajiki et al. (2007)Korea 61 64 <3–136 Lim et al. (2009a)Belgium 21 100 0.2–169 Geens et al. (2010)

Beverage cans (ng/mL)Spain 11 64 <0.05–0.61 Gallart-Ayala et al. (2010)Canada 69 100 0.03–4.5 Cao et al. (2009a)Belgium 45 91 <0.02–8.1 Geens et al. (2010)Portugal 30 70 <0.01–4.7 Cunha et al. (2011)


are often used as internal coatings of cans to prevent direct contactbetween the metal can walls and the food or beverage, and to pro-tect the cans from rusting and corrosion (Cao et al., 2011; Goodsonet al., 2002). These protective coatings are also used on metal lidsfor foods in glass jars (Cao et al., 2011). Due to an incomplete poly-merization process, residues of BPA monomer in PC containers andcoatings can migrate into foods, especially during storage and pro-cessing at elevated temperatures (Cao et al., 2011; Geens et al.,2010; Noonan et al., 2011).

2.1. Epoxy resins

2.1.1. MigrationThe influence of damage, storage conditions and heating on the

migration of BPA was studied by Goodson et al. (2004). Emptyepoxy phenolic coated cans were filled with four foods and 10%ethanol as food simulant. Filled cans of each food type or simulantwere sealed and processed using usual conditions. Cans werestored at 5, 20 or 30 �C and analyzed at different time intervals(up to 9 months). Half of the cans were dented in order to evaluatethe effect of damage on the migration. Between 80–100% of freeBPA already present as free monomer in the coating had migratedinto the food during sterilization. Extended storage at various tem-peratures or damaging of the can did not change the migrated BPAlevels (Goodson et al., 2004).

The effect of heat treatment on the migration of BPA was ob-served by Munguia-Lopez et al. (2002, 2005) and Munguia-Lopezand Soto-Valdez (2001). Most of the BPA migrated during heattreatment (121 �C and 90 min) using an aqueous food simulant, afatty food simulant, jalapeño peppers or tuna fish (Munguia-Lopezand Soto-Valdez, 2001; Munguia-Lopez et al., 2005). For jalapeñopeppers, which are more acidic than tuna, sterilization for 9 minat 100 �C had a minimal effect on the migration of BPA, both forthe aqueous food stimulant and the acid food simulant. Due tothe milder heat processing conditions for jalapeño pepperscompared to tuna, part of the residual BPA remained on the coatingafter processing. Afterwards, BPA increased during storage time,especially during the first 40 days (Munguia-Lopez and Soto-Valdez, 2001; Munguia-Lopez et al., 2002). Kang and Kondo(2003) reported that the temperature has more influence on themigration of BPA from an epoxy can coating in water than the heat-ing time.

2.1.2. LevelsThe dominant contribution of canned food to the overall expo-

sure to BPA was confirmed in several intervention studies. In astudy of Carwile et al. (2011), the urine of 75 volunteers who con-sumed one serving of canned soup during five days showed a spec-tacular increase of 1200% in urinary BPA concentrations comparedto urine concentrations following the consumption of fresh food

during five days. Braun et al. (2011a) observed higher BPA concen-trations in urine of pregnant women who consumed at least once aday canned vegetables compared to those who did not consumecanned vegetables. In a dietary intervention study where volun-teers were subjected to a 3-day ‘‘fresh food’’ diet that was notcanned or packaged in plastic, Rudel et al. (2011) observed a 66%decrease in urinary BPA concentration compared to the concentra-tions prior to the intervention.

Several studies worldwide determined BPA in canned food,including US (Noonan et al., 2011; Schecter et al., 2010), Canada(Cao et al., 2010, 2011), Japan (Sajiki et al., 2007), Korea (Limet al., 2009a), New Zealand (Thomson and Grounds, 2005), UK(Goodson et al., 2002) and Belgium (Geens et al., 2010). The samplesize, detection frequency and concentration range are summarizedin Table 1.

In all studies, large variations in BPA concentrations were foundbetween different products of the same food type, but also be-tween different lots of the same product. Noonan et al. (2011) ob-served a 100-fold difference (2.6–310 ng/g) between the minimaland maximal BPA values in peas, while green beans had a 30-folddifference (22–730 ng/g) between brands. Geens et al. (2010) ob-served also a large variation (1.2–82 ng/g) between five brands ofcorn. While some studies reported tuna fish to have the highestcontamination of BPA (Cao et al., 2010; Lim et al., 2009a), in otherstudies, tuna samples had the lowest concentrations of BPA (Noo-nan et al., 2011). Such variation is probably due to the differentproprietary composition of the coatings from can manufacturersand to the different can styles or coating choices for various prod-ucts used by the food producers (Noonan et al., 2011). Unfortu-nately, these differences have been less investigated and are notsubject of any regulation. In contrast, the lot-to-lot variability forsamples of the same food type and brand was smaller than the var-iability between and within foods (Noonan et al., 2011). In foodwhere both a solid portion and liquid supernatant are present,BPA intends to partition into the solid part (Geens et al., 2010; Noo-nan et al., 2011). Yet, the BPA concentration in the solid partseemed to be dependent on the type of food. While for corn (Yos-hida et al., 2001), green beans and peas (Noonan et al., 2011), BPAwas partitioned in the solid part of the food, BPA remained in theaqueous solution for peeled oranges (Yoshida et al., 2001). It isnot clear whether the migration of BPA into the solid portion couldbe explained by the absorption to fibers, by the fat content of thefood or by other mechanisms (Yoshida et al., 2001).

Similar to food cans, BPA can also migrate from beverage cans.The most relevant studies are summarized in Table 1. In contrast tocanned food samples, BPA concentrations in canned beveragesshowed a more narrow range. For the Canadian and the Belgianstudy, respectively 85% and 75% of the samples had concentrationsbelow 1 ng/mL (Cao et al., 2009a; Geens et al., 2010). The lowerconcentrations found in beverages can possibly be explained by


the differences in the can type, can coating and sterilization condi-tions between food and beverages (Geens et al., 2010). Besides BPA,Cunha et al. (2011) detected BPB in 50% of the canned beverages(range 0.07–0.16 ng/mL). Gallart-Ayala et al. (2010)) could not de-tect BPB, Bisphenol E or BPS in any soft drinks, but they could de-tect BPF in two samples (0.14 and 0.22 ng/mL).

Next to the use of epoxy resins as protective layer of food andbeverage cans, epoxy resins can also be used as internal coatingon metal lids for food in glass jars, a source scarcely investigateduntil now. Whilst the contact between the lid and the food is rarewhen compared to contact between food and can, such contact cansometimes occur. It is caused by transportation of the cans, byshaking, as well as by accidental storage in a non-vertical position(Cao et al., 2009b). Therefore, Cao et al. (2009b) determined BPA in99 baby food products of seven Canadian brands in glass jars withmetal lids. BPA was detected in 85 samples (86%), from which 69samples (70%) had levels of less than 1 ng/g with an overall averageconcentration of 1.1 ng/g.

Cao et al. (2011) investigated 154 food composite samples fromthe 2008 total diet study in Quebec City, Canada. BPA was detectedin 55 of the 154 food samples (36%) tested. High concentrations ofBPA were found mostly in the composite canned foods, with thehighest BPA level being observed in canned fish (106 ng/g).

BPA was also detected in some foods that are nor canned, norpackaged in jars, such as yeast, baking powder, cheese, bread, cere-als, and fast foods. The source of BPA in these food items was sug-gested to be the packaging paper, especially plastic packaging filmin PVC, or BPA could be introduced during the production process ifequipments or containers with epoxy coating or plastic parts havebeen used. BPA contamination of fast food could be due to thewrapping paper or BPA may already have been present in theingredients used to prepare the fast food. BPA intakes from 19out of 55 samples in which BPA was detected, accounted for morethan 95% of the total dietary estimated intake in Canada, and mostof the 19 samples were either canned or in jars. The remaining 36samples in which BPA was detected, contributed with only 5% tothe estimated dietary intake. Therefore, the intake of BPA fromnon-canned foods was estimated to be low (Cao et al., 2011).

2.2. Polycarbonate

2.2.1. Migration and hydrolysisOn January 28, 2011 the European Commission (EC) published a

Directive that PC may not be used any longer for baby bottles (EC,2011b). Additionally the EC issued the regulation No. 321/2011(EC, 2011c), indicating that baby bottles made of PC may not beproduced any more from the 1st of May 2011 and not be put onthe market from the 1st of June 2011. Although similar bans onthe production, import and sale of PC baby bottles have been intro-duced in Canada and several US states, exposure through PC can

Table 2Migration of BPA from polycarbonate baby bottles. BPA has a specific migration limit of 6

Reference Highest BPAconcentration (lg/L)

Releva

De Coensel et al. (2009) 0.30 60 s aEhlert et al. (2008) 0.73 3 cycleLe et al. (2008) 1.33

7.677 days

Kubwabo et al. (2009) 6.5 MigratMaragou et al. (2008) 14.3 20 cyc

tempeNam et al. (2010) 18.5 100 timBiedermann-Brem and Grob, 2009 137 PrevioBiedermann-Brem et al. (2008) �500 A slan

run ofCao and Corriveau (2008b) 521 Heatin

still be of relevant in other countries and by the use of ‘‘old’’ PCbaby bottles or other PC food contact applications.

BPA can leach from PC into liquids through two different pro-cesses: diffusion of residual BPA present in PC after manufacturingand hydrolysis/aminolysis of the polymer (Aschberger et al., 2010).Experiments using official simulants usually report the migrationof BPA which is, even under rather drastic conditions (such as 1 hat 100 �C), typically in the range of 0.1–1 lg/L (Biedermann-Bremand Grob, 2009). For a usual migration behavior, a decrease is ob-served after continued use. The low migration from PC baby bottlesinto food simulants was confirmed in several recent studies (Bie-dermann-Brem et al., 2008; Santillana et al., 2011; Simoneauet al., 2011). Most of the baby bottles showed migration belowthe detection limit of 0.1 lg/kg (Simoneau et al., 2011) or<0.4 lg/L from new baby bottles and after 30 washing cycles (Bie-dermann-Brem et al., 2008).

Increased migration of BPA from PC baby bottles was observedfor higher temperatures and longer testing periods (Biedermann-Brem and Grob, 2009; De Coensel et al., 2009; Kubwabo et al.,2009; Le et al., 2008; Lim et al., 2009b; Nam et al., 2010). An in-crease in the BPA migration rate up to 55-fold during exposure ofthe PC to boiling water (100 �C) compared to water at 20 �C wasobserved (Le et al., 2008). Microwave heating did not seem to havean effect, and migration was mainly temperature dependent (Ehl-ert et al., 2008; De Coensel et al., 2009).

Contrary to the usual migration behavior, where a decrease inmigration or a constant migration was observed after repeateduse of the PC bottles, several studies reported an increase in theBPA migration over time, due to hydrolysis of the PC (Brede et al.,2003). Biedermann-Brem and Grob (2009) revealed that the higherconcentrations can be due to aging that increases the wettability ofthe bottle wall, which in turn promotes the adherence of water tothe bottle wall. Drying in the dish washing machine causes dis-solved salts to reconcentrate on the bottle wall and to be baked ontothe PC at elevated temperature. They may promote the degradationof the polymer and the release of BPA, especially when alkali chem-icals are deposited, such as washing solutions (Biedermann-Bremand Grob, 2009). Rinsing of bottles before the drying step couldovercome this ‘‘baking’’ and, thus, the release of high BPA concen-trations. However, preparing a drink according to the usual recom-mendations results usually in a BPA release <0.5 lg/L (Biedermann-Brem and Grob, 2009). Similarly, aminolysis of PC was observedafter contact with two biogenic amines (1,4-diaminobutane and tri-methylamine) (Maia et al., 2010) or after contact with alkalinedetergent solutions (Maia et al., 2009).

The highest migration or release of BPA from PC bottles wasobserved under conditions which are not likely to occur under nor-mal use, i.e. at elevated temperature or contact time (Table 2). DeCoensel et al. (2009) reported only very low migration levels ofBPA (6–13 ng/L) when the bottles are used under normal conditions

00 lg/kg (EC, 2011a).

nt conditions

nd 1000 W (65 �C)s of 100 �C in microwave oven (�3 min)at room temperature 24 h at 100 �C

ion in water (24 h at 60 �C)les of cleaning-sterilization-filling with boiling water and left at roomrature for 45 min

es for 30 min in steam bath at 95 �Cusly boiled tap water (pH 9.5) in microwave for 10 min Release of BPAted position of the bottle in the dishwasher, hindering the detergent solution tof and rinsing before dryingg water at 70 �C for 6 days

Table 3Estimated intake of BPA in children and adults.

Age category Estimation through dietaryexposure (lg/kg bw/day)

ChildrenEFSA (2006) Infants (3–12 month)

Children0.2–135.3

Health Canada (2008) 1–4 years5–11 years

0.26–1.980.15–1.28

Chapin et al. (2008) Infants-bottle fedInfants-breast fedChildren (6–12 m)Children (2–6 years)

1–110.2–11.7–130.04–14.7

FDA (2009) 0–12 m12–24 m>2 years

0.3–0.60.5–1.10.1–0.3

ANSES (2010) Infants (<36 m)Children (3–17 years)

0.1–0.50.2–0.6

WHO (2010) Infants 0–6 mInfants 6–36 mChildren > 3 years

0.01–4.50.01–3.00.2–1.9

AdultsEFSA (2006) Adults 1.5Health Canada (2008) 12–19 years

>20 years0.09–0.730.07–0.60

Chapin et al. (2008) Adults 0.008–1.5FDA (2009) >2 years 0.1–0.3ANSES (2010) Adults 0.1–0.3WHO (2010) Adults 0.4–4.2


(20 s at 1000 W in the microwave oven at 37 �C). During normaluse, the released BPA quantities are negligible (maximum 2 ngper feeding) and far below the TDI value.

2.2.2. LevelsThe effect of migration of BPA from PC drinking bottles was

illustrated in an intervention study where volunteers were re-quested to consume all cold beverages from PC drinking bottlesduring one week. An increase of 69% in urinary BPA concentrationswas observed after one week compared with urinary levels ob-tained after a wash-out period of one week, where no use of PC-bottles was allowed (Carwile et al., 2009).

In Canada, Cao and Corriveau (2008a) could not detect BPA in 51non-PC bottled water products (detection limit 0.5 ng/mL). How-ever, BPA was detected in 4 out of 5 bottled water in PC products(<0.5–1.4 ng/mL). In a 5-week experiment, levels of 8.8 and6.5 ng/mL were measured in two bottles. Therefore the authorswarn for higher BPA levels that could be detected in some PC bot-tled water products due to the accidental or careless exposure toheat (e.g. sun) for extended periods of time during storage andtransportation (Cao and Corriveau, 2008a). In a Greek study (Amir-idou and Voutsa, 2011), BPA was determined in water from fivePET-bottles with a median concentration of 4.6 ng/L. Water fromone PC bottle contained 112 ng/L which increased to 170 ng/L after30 days of sun exposure. The maximum daily intake through bot-tled water, assuming a daily intake of 2 L water was estimated tobe merely 0.006 lg/kg bw/day.

PC is used also for water pipes and epoxy-phenolic resins arewidely used as a surface-coating on residential drinking waterstorage tanks (Bae et al., 2002). Li et al. (2010) detected BPA intap water from six different drinking water plants in Guangzhou,China in concentrations between 15 and 317 ng/L. Daily mean in-take of BPA of adults was estimated to be 148 ng/day from drinking2 L of tap water. Yet, more data are needed to quantify the possibledietary exposure to BPA via drinking water (EFSA, 2006).

2.3. Other food contact applications

BPA is rarely measured in non-canned foods, thus, the contribu-tion from the non-canned foods to the overall dietary intake of BPAis not well known (Cao et al., 2011). Geens et al. (2010) determinedBPA in 16 solid food samples packaged in glass, plastic, paper andlaminated paperboard/polyethylene carton (Tetra Pak). BPA couldbe detected in all food samples in a concentration range of 0.1–1.28 ng/g with an average concentration of 0.46 ng/g. This meanconcentration is about 100 times lower than the average concen-tration in similar food types, packaged in cans which were exam-ined in the same study. BPA could not be detected above thequantification limit of 0.02 ng/mL in five beverages packaged inPET and Tetra Pak (Geens et al., 2010). Also Sajiki et al. (2007)found considerably lower concentrations of BPA in 15 out of 23food samples (range <1–14 ng/g) packaged in plastic and in 4 outof 16 food samples (<0.2–1 ng/g) packaged in paper, comparedwith the food samples packaged in cans.

No BPA was observed to migrate from EcoCare™ lined alumin-ium, stainless steel, or Tritan™ plastic water bottles during anincubation period of 120 h (detection limit 0.05 ng/mL). In con-trast, detectable amount of BPA were leached from PC bottlesand epoxy-lined aluminum bottles (Cooper et al., 2011).

BPA was found to be present in commercial PVC cling films andplastic sheeting bags available on the market in Spain and migra-tion studies suggested it would migrate into food (Lopez-Cervantesand Paseiro-Losada, 2003). Yet, the former application of BPA in thePVC polymerisation process by some EU manufacturers appearedto have ceased (EFSA, 2006). Therefore, based on this information,no BPA exposure from food contact uses of PVC should be expected

in the EU today, however PVC materials which were produced priorto this action may still be in use.

2.4. Intake estimation from food exposure

The estimated intake through food by different national andinternational agencies is summarized in Table 3. These estimationsare sometimes based on highest observed concentrations or migra-tion values or are derived using 95th percentile estimates of con-sumption. The highest estimated BPA dietary exposures were for0–6 months of age infants who were exclusively fed on canned li-quid infant formula using PC bottles. In this case, sources of BPAexposure include migration from both the formula packaging andfrom the PC bottle (WHO, 2010). However, in all studies, eventhe worst case, estimates stay below the current TDI.

Mean exposures for infants fed with infant formula using PCbottles were 2.0–2.4 lg/kgbwper day, with 95th percentile expo-sures ranging from 2.7 to 4.5 lg/kgbwper day (WHO, 2010). In-fants who were either fed with formula from non-PC bottles orexclusively breastfed had substantially lower estimated meanBPA exposures (0.01 lg/kgbwper day from powdered formula,0.5 lg/kgbwper day from canned liquid formula and 0.3 lg/kgbw per day from breast milk), compared to those exclusively fedon infant formula using PC bottles. Once solid foods are introduced(at 6–36 months), exposure to BPA decreases relative to bodyweight.

For children above 3 years, the highest mean BPA exposure wasestimated to be 0.7 lg/kgbwper day, with a maximum up to1.9 lg/kgbwper day (Table 3). Depending on the extent of pack-aged food (canned) in the diet, adult BPA exposures were compara-ble to those for children above 3 years: a highest mean exposure of1.4 lg/kg bw per day, with a maximum exposure up to 4.2 lg/kgbwper day (Table 3). It was assumed that all exposure to BPAfrom the diet was in the form of unconjugated BPA. These calcu-lated international dietary exposure estimates (WHO, 2010) areconsistent, but slightly higher than those obtained using data re-ported from comparable national surveys.

In Canada, dietary intake estimates of BPA by different age-sexgroups were made based on the concentrations found in the food


composites combined with data of the 24-h diet recall from theNutrition Canada Survey (Cao et al., 2011). Dietary intakes of BPAwere low for all age–sex groups, with 0.17–0.33 lg/kgbw/day forinfants, 0.082–0.23 lg/kgbw/day for children aged from 1 to19 years, and 0.052–0.081 lg/kgbw/day for adults. Where Caoet al. (2011) included both canned and non-canned food in theirestimation, other studies made intake estimations only based oncanned food and only for adults. For example, Thomson andGrounds (2005) estimated an intake of 0.008 lg/kgbw/day inNew Zealand, Geens et al. (2010) found 0.015 lg/kgbw/day in Bel-gium, while Lim et al. (2009a) estimated an intake of 0.030 lg/kgbw/day in Korea. Mariscal-Arcas et al. (2009) included, next tocanned food, also migration from polycarbonate tableware andestimated an intake of 0.030 lg/kgbw/day.

Overall, intake of BPA from food is well below the current TDI of50 lg/kgbw/day. However, more knowledge is necessary on the ef-fect of food processing, preparation and cooking procedures on BPAlevels in the final cooked foods. Since PC tools and containers withepoxy coatings may be used during food preparation for cooking,BPA could be introduced into the final cooked foods due to migra-tion from PC and coatings (Cao et al., 2011).

3. Non-food sources to bisphenol-A

3.1. Dust

Because of the low vapor pressure of BPA and, therefore, its lowconcentrations in air, inhalation of BPA from air is unlikely to be animportant exposure source (Dekant and Völkel, 2008).

Ingestion of house dust has been demonstrated to be an impor-tant exposure pathway to several contaminants in young childrendue to their more frequent hand-to-mouth contact and larger in-take of dust compared to adults (Jones-Otazo et al., 2005; Calafatet al., 2008). Due to the wide use of BPA in a variety of indoor appli-cations and consumer products, such as epoxy-based floorings,adhesives, paints, electronic equipments, and printed circuitboards, volatilization and/or leaching of BPA from these productsare a source of contamination of indoor dust (Loganathan and Kan-nan, 2011). Consequently, BPA was detected in indoor dust with ahigh detection frequency and ranged widely up to �10,000 ng/gdust (Geens et al., 2009). Median concentrations of BPA in variousstudies ranged between 422 and 1460 ng/g in US (Völkel et al.,2008; Geens et al., 2009; Loganathan and Kannan, 2011).

Higher concentrations were observed in laboratories (Logana-than and Kannan, 2011) and offices (Geens et al., 2009) most prob-ably due to the use of more electric and electronic equipment andfurniture than in homes. Contrary, lower concentrations were ob-served in dust samples from daycare centers in US (Rudel et al.,2003; Wilson et al., 2007). Toddlers have a more frequenthand-to-mouth contact and will therefore have a higher dust in-take. Although the amount of dust daily ingested is uncertain,the intake of BPA from dust ingestion is low and was estimated

Fig. 2. Structure of thermal pape

to be less than 0.006 lg/kg bw/day for toddlers and less than0.0005 lg/kg bw/day for adults (Geens et al., 2009; Loganathanand Kannan, 2011). The contribution of dust to the total intake ofBPA is therefore probably less than 1–5%.

3.2. Thermal paper

BPA is used as an additive in thermal paper made for printersrelying on the thermal transfer technology, whereby BPA is usedas a color developer. In these papers, one side is coated with a pow-dery layer of BPA (Lassen et al., 2011). Under heat or pressure, BPAreacts with the thermal paper dye to produce a color-developingcomplex (Fig. 2). This technique is mainly used in lightweightprinting devices, such as cash registers or credit card terminals.

Many people come in contact with thermal paper on a daily ba-sis. The presence of BPA in thermal paper may contribute to theoverall exposure by oral intake (direct contact of unwashed handswith food or mouth) or by dermal exposure. Moreover, thermal pa-per is also a major source of contamination of recycled paper withBPA (Takahashi et al., 2002; Zalko et al., 2011). Braun et al. (2011a)already reported the higher levels of urinary BPA of cashiers, whichmight have a higher skin contact with BPA-containing thermal pa-per compared to the general population. Worldwide, BPA was de-tected in thermal paper (Denmark, Sweden, Switzerland, US)with a detection frequency between 44% and 100%. BPA concentra-tions in the thermal paper were up to 2.3% (Biedermann et al.,2010; EWG, 2010; Lassen et al., 2011; Liao and Kannan, 2011a;Mendum et al., 2011; Östberg and Noaksson, 2010) (Table 4). Liaoand Kannan (2011a) could not detect BPA in all seven thermal pa-pers from Japan, most probably due to the phase-out of BPA inthermal paper in Japan in 2001.

The amount of BPA transferred to the skin after holding such apaper for 5 s was between 0.2 and 6 lg BPA with an average of1.1 lg per finger (Biedermann et al., 2010). If the fingers werewet or very greasy, the transferred amount was about 10 timeshigher. Repeated contact with fresh recorder paper did not give asignificant increase in BPA on the skin, indicating equilibrium be-tween the BPA concentration in the paper and on the surface layerof the skin. Biedermann et al. (2010) could not conclude whetherBPA passed through the skin, but found that BPA can enter the skinto such a depth that it can no longer be washed off. For normalskin, a potential exposure of 71 lg/day was estimated when touch-ing the most contaminated paper frequently during a working dayof 10 h (Biedermann et al., 2010). Mielke et al. (2011) predictedthat dermal exposure can have a relevant contribution to the totalBPA exposure.

Based on the worst case dermal exposure of 71 lg/day (0.97 lg/kgbw/day) determined by Biedermann et al. (2010), and on the ex-tent of dermal absorption recently published, (10% (EU, 2008), 13%(Mørck et al., 2010), 46% (Zalko et al., 2011) and 60% (Biedermannet al., 2010)), dermal exposure can result in an uptake between7.1 lg/day (0.1 lg/kgbw/day) and 42.6 lg/day (0.58 lg/kgbw/

r (from Lassen et al., 2011).

Table 4Overview of BPA in thermal paper.

Country Sample size Detection freq. (%) % (g BPA/100 g paper) Refs.

Denmark 12 65 n.d–1.7 Lassen et al. (2011)Sweden 16 100 0.6–2.3 Östberg and Noaksson (2010)Switzerland 13 85 <5.10�5–1.7 Biedermann et al. (2010)US 36 44 0.8–2.8 EWG (2010)US, Boston 10 80 <0.09–1.7 Mendum et al. (2011)US, Japan, Korea, Vietnam 103 94 <1.10�7–1.4 Liao and Kannan (2011a)


day). Similarly, a Danish study reported a realistic worst case sce-nario which resulted in a daily uptake of 240 lg BPA (Lassen et al.,2011). In this scenario, it is assumed that the receipts are touchedwith humid fingers and that 50% of the quantity left on the skin isabsorbed (Lassen et al., 2011). However, the actual exposure of thegeneral consumer will mostly be lower.

3.3. Other types of papers

Thermal paper can also be the primary cause of the contamina-tion of paper currencies. Paper currencies from 21 countries wereanalyzed for BPA (Liao and Kannan, 2011b). BPA was found in all pa-per currencies at concentrations ranging up to 82.7 lg/g. The con-tamination of the paper currencies can probably be explained byfrequent contact with thermal paper in a wallet. Because the BPAused in thermal paper is not covalently bound, it can be easily trans-ferred from thermal receipt papers to other objects, including papercurrencies. BPA may be also present also in the production processof currency paper. The estimated daily intake of BPA through der-mal absorption from handling paper currencies was on the orderof a few nano grams per day (Liao and Kannan, 2011b).

It has been estimated that approximately 30% of thermal papersenter recycling streams of municipal wastepaper. Recycling ofthermal paper can introduce BPA into the paper production cycle(Liao and Kannan, 2011a). Vinggaard et al. (2000) showed that,while virgin paper contained no or negligible amounts of BPA, lev-els in the recycled paper ranged from 0.6 to 24 lg BPA/g of kitchenroll. Similarly, a Japanese study examined paperboard and papersused for food packaging. In the virgin paper and paperboard, con-centrations between (0.034–0.36 lg/g) were detected, while theconcentrations in recycled paper and paperboard were >10-foldhigher (range 0.19–26 lg/g) (Ozaki et al., 2004). More than 80%of others papers, including flyers, tickets, newspapers, toilet paper,contained BPA in concentrations ranging up to 14.4 lg/g. Thus, BPAconcentrations in ‘‘other’’ papers were 3–4 orders of magnitudelower than in thermal paper, most probably due to the recyclingof thermal paper (Liao and Kannan, 2011a). The exposure to BPAfrom other papers will have an insignificant contribution to theoverall exposure. Liao and Kannan (2011a) made an assessmentthrough dermal exposure of BPA from thermal and ‘‘other’’ paper.The median dermal exposure to BPA of the general populationwas 17.4 ng/day, while this was 1303 ng/day for the occupationallyexposed population. Thermal paper contributed for more than 98%to this value. Liao and Kannan (2011a) calculated that for an over-all exposure to BPA of 1 lg/kg bw/day, paper could contribute 1.6–51% in an occupationally exposed population.

3.4. Dental materials

Dental composite resins consist of a mixture of co-monomersand are most commonly based on bisphenol-A glycidyl methacry-late (bis-GMA). In addition to bis-GMA, these resins contain othermonomers to modify the properties, e.g. bisphenol-A dimethacry-late (bis-DMA). Although BPA is not used itself in composite resins,

it might be present as an impurity from the synthesis process (Fle-isch et al., 2010; Fung et al., 2000; Nathanson et al., 1997; VanLanduyt et al., 2011). BPA can also leach into the saliva as a resultof bis-DMA hydrolysis through esterases present in the saliva (re-viewed by Van Landuyt et al., 2011).

Several in vivo studies measured BPA in saliva after sealantplacement. Salivary BPA levels decreased over time; the highestexposures were measured immediately after sealant placement.BPA exposure after sealant placement is most likely an acute event,yet none of the studies could detect BPA 3 h after sealant place-ment. Possibly, analytical methods used in these studies werenot sensitive enough to detect extremely low doses of BPA thatchronically leach from the resin over longer periods of time. Hence,chronic low-dose BPA exposure after dental sealant placementcannot be ruled out (Fleisch et al., 2010).

The relevance of the released amounts of BPA from dental mate-rials in vitro has recently been reviewed in a meta-analysis done byVan Landuyt et al. (2011). It was computed that one full crown res-toration of a molar may release 13 lg BPA in the average case sce-nario or 30 mg BPA in the worst case scenario, both after 24 h. Theaverage BPA release (0.2 lg/kg body weight/day for a personweighting 60 kg) is 250-fold lower than the TDI of 50 lg/kg bodyweight/day, but 10-fold higher than the TDI in the worst case sce-nario. This indicates that the 24-h release of BPA from dental mate-rials is relevant in patients with multiple or large restorations andthat resin-based dental materials may represent a relevant sourceof BPA in such patients (Van Landuyt et al., 2011). Sealants pro-duced by different manufacturers released markedly differentamounts of BPA (Vandenberg et al., 2007).

Von Goetz et al. (2010) estimated the chronic exposure afterdental surgery to be 215 ng BPA/day. This estimation was basedon the measurement of 0.3 ng/mL in the saliva of one out of 21individuals at 120 h after surgery. It probably represents a worst-case scenario for chronic exposure, since concentrations in salivawill decrease further over time and only one individual had stillmeasurable concentrations after 120 h.

3.5. Medical devices and healthcare applications

A small fraction of the BPA-based polymers polycarbonate andpolysulfone is used in medical and healthcare applications suchas PC eye lenses, tube connections, blood oxygenators, inhalerhousing, and newborn incubators, as well as polysulfone surgicaltrays, nebulizers, and humidifiers (Geens et al., 2011). BPA can alsoleach into a drug formulation which most likely occurs with liquidand suspension formulations that are packaged in PC container-closures or metal canisters with epoxy lining (FDA, 2009).

PVC, which may also contain BPA, is used in the manufacturingof medical products, such as those found in the neonatal intensivecare units, including bags containing intravenous fluids and totalparenteral nutrition and tubing associated with their administra-tion; nasogastric and enteral feeding tubes; and umbilicalcatheters. In a study of Calafat et al. (2009), BPA was analysed inurine from 42 low-birth-weight infants in neonatal intensive care


units using a large number of PVC-containing devices, such asmechanical and high-frequency ventilation, surgery, and cardiaccatherization. Median concentrations of BPA in these premature in-fants were one order of magnitude higher than the median concen-tration and almost twice the 95th percentile of the generalpopulation (children 6–11years who were examined as part ofthe NHANES 2003–2004) (Calafat et al., 2009).

Hemodialysis patients can be exposed to substantial amounts ofBPA due to the use of PC as casing and the hollow-fibers’ hemodi-alysis membrane often made of polysulfone. Moreover, the re-leased BPA is directly introduced into the blood circulation.Although not an exposure source for the general population,hemodialysis may be an important contributor for this specificgroup (Geens et al., 2011; Haishima et al., 2001; Yamasaki et al.,2001). Almost no data exist to quantify the dose of BPA that treatedpatients receive; further research is therefore highly necessary(FDA, 2009).

3.6. Other non-food sources

In a Danish study, the migration from the shield and ring ofbaby dummies was examined. These parts can be made of PC,although it has been largely replaced by polypropylene and co-polyester. Even when the shield and ring contained PC, migrationof BPA into sweat and saliva was low and the calculated exposureto BPA in dummies was far below the BPA exposure from baby bot-tles (Lassen et al., 2011).

4. Toxicokinetics and metabolism of bisphenol-A

The toxicokinetics of BPA has been studied in rodents, non-hu-man primate and humans (Doerge et al., 2010a,b; Völkel et al.,2002, 2005). After oral administration, BPA undergoes a rapid firstpass metabolism in the intestine and liver, being completely ab-sorbed from the gastrointestinal tract. BPA is not extensivelymetabolized via Phase I reactions, but it is rapidly conjugated withglucuronic acid (Phase II metabolism) to the non-active BPA-glucu-ronide in the gut wall and liver. Minor amounts of BPA might alsoreact with sulfate to form BPA-sulfate. The formation of BPA conju-gates is considered a detoxification process (Matthews et al., 2001;Snyder et al., 2000) and only the free BPA forms display estrogenicactivity (Matthews et al., 2001). The BPA conjugates formed in theliver are delivered to the blood in humans to reach the kidney,being further excreted in the urine with terminal half-lives of lessthan 6 h (Völkel et al., 2002, 2005). The applied doses were com-pletely recovered in urine; hence, BPA exposure can be estimatedfrom urinary levels (Völkel et al., 2002). BPA ingested by inhalationor dermal contact does not undergo first pass effect and will there-fore be eliminated at a slower rate.

In adult rhesus monkeys, the concentration–time profile afteroral administration of BPA was remarkably similar to humans, gi-ven a similar dose (Doerge et al., 2010b). Minimal pharmacokineticdifferences were observed between neonatal and adult monkeysfor the free form of BPA, which was present in less than 1% ofthe total circulating concentration of BPA (Doerge et al., 2010b).In rodents, BPA-glucuronide is subject to enterohepatic recircula-tion, which prolongs elimination processes, thereby increasinginternal exposures to BPA, and leads to extensive fecal excretion(Pottenger et al., 2000). The absence of enterohepatic circulationof BPA-glucuronide in humans is most likely due to a higherthreshold for biliary elimination as compared to rats.

Several tissues, including human liver and kidney, contain b-glucuronidase in membranes of lysosomes and the endoplasmicreticulum (Sperker et al., 1997). It has been suggested that b-glucu-ronidase activity in tissues, especially placenta, could reverse the

detoxification of BPA at the tissue level (Ginsberg and Rice,2009). The experimental evidence to support this hypothesis is lar-gely indirect and inconsistent with the rapid elimination of agly-cone BPA from the circulation in adult non-human primates andhumans (Völkel et al., 2002). Also viable human skin explants effi-ciently absorbs and metabolizes BPA. About 46% of the applieddose of BPA was absorbed and largely transferred into BPA-glucu-ronide and BPA-sulfate (Zalko et al., 2011).

5. Human biomonitoring

As a non-persistent chemical with an elimination half-life of afew hours, the BPA concentrations in blood are lower than thosein urine and decrease quickly after the exposure (Needham andSexton, 2000). As a result, BPA will be non-detectable in a largerproportion of blood samples with the current analytical technology(WHO, 2010). Moreover, it is difficult to rule out contaminationwith trace levels of free BPA during sample collection, storageand analysis because of the ubiquitous presence of BPA in the envi-ronment (WHO, 2010; Markham et al., 2010; Völkel et al., 2008).Even detectable concentrations do not thus necessarily reflectBPA exposures.

Since BPA is rapidly and almost completely excreted as BPA-con-jugates, urine is the matrix of choice for biomonitoring. Long-termdaily intake of BPA leads to steady-state BPA concentrations in theng/mL range in human samples (Welshons et al., 2006). Urinaryconcentrations of total (free plus conjugated) BPA have often beenused to evaluate exposure to BPA from all sources (Vandenberget al., 2010). Several biomonitoring studies have been conductedin North America, Europe and Asia, revealing the worldwide expo-sure to BPA. The most important studies are summarized in Table 5.

A study documenting measurable urinary BPA levels in Mexicanwomen provides preliminary evidence that pregnant women whodelivered prematurely (<37 weeks gestation) had higher urinaryconcentrations of BPA compared to women delivering after37 weeks (Cantonwine et al., 2010). The impact of gestational ver-sus childhood BPA exposures is unclear. In a recent US study, ges-tational BPA exposure affected behavioral and emotionalregulation domains at 3 years, especially among girls. These resultssuggested that gestational BPA exposure might be associated withanxious, depressive, and hyperactive behaviors related to impairedbehavioral regulation at 3 years (Braun et al., 2011b).

Two recent large-scale studies which included 2514 and 5476participants were performed in the USA and Canada, respectively.Exposure to BPA was ubiquitous with a detection frequency ofmore than 90% in both studies (Calafat et al., 2008; Bushniket al., 2010). Also in seven Asian countries, BPA was detected in94% of the samples (Zhang et al., 2011). In the US study, highesturinary concentrations were detected in adolescents (12–19 years)followed by children (6–11 years) and adults (>19 years). Afteradjusting BPA levels for creatinine, children had the highest BPAconcentrations, followed by adolescents and adults (Calafat et al.,2008). Also in the Canadian study (Bushnik et al., 2010), creatinineadjusted BPA levels were higher in the youngest age category (6–11years) than for the other age categories. In the GerES IV studyin Germany, children in the age category 3–5years had higher con-centrations than the 6–8years; 9–11years; and 12–14years agecategory (Becker et al., 2009). Vandenberg et al. (2010) also con-cluded that there is an indication that young children are submit-ted to the highest exposure risk.

For practical reasons, biomonitoring studies with urine samplesgenerally collect single spot urine samples instead of 24 h urinesamples. Because of BPA’s short elimination half-life, spot urinesamples primarily reflect the exposure that occurred within a rel-atively short period before urine collection (Koch and Calafat,

Table 5Overview of the most recent worldwide biomonitoring studies in urine.

Country Population Concentrations Exposure Det. Freq. (%) Refs.

US 2514 (P6–P60 years)314 (6–11 years)713 (12–19 years)950 (20–59 years)537 (P60 years)

GM 2.6 ng/mL (2.6 lg/g cr)GM 3.6 ng/mL (4.3 lg/g cr)GM 3.7 ng/mL (2.8 lg/g cr)GM 2.6 ng/mL (2.4 lg/g cr)GM 1.9 ng/mL (2.3 lg/g cr)

GM 0.047 lg/kg bw/dayGM 0.065 lg/kg bw/dayGM 0.071 lg/kg bw/dayGM 0.053 lg/kg bw/day (20–39 years)GM 0.038 lg/kg bw/day (40–59 years)GM 0.034 lg/kg bw/day

93 Calafat et al.(2008) Lakind andNaiman (2008)

US 394 adults GM 1.33 ng/mL (1.36 lg/g cr) GM 0.023 lg/kg bw/daya 95 Calafat et al.(2005)

Canada 5476 6–79 years6–11 years12–19 years20–39 years40–59 years60–79 years

GM 1.16 ng/mL (1.40 lg/g cr)GM 1.30 ng/mL (2.00 lg/g cr)GM 1.50 ng/mL (1.31 lg/g cr)GM 1.33 ng/mL (1.49 lg/g cr)GM 1.04 ng/mL (1.33 lg/g cr)GM 0.90 ng/mL (1.26 lg/g cr)

GM 0.025 lg/kg bw/dayGM 0.031 lg/kg bw/dayGM 0.026 lg/kg bw/dayGM 0.020 lg/kg bw/dayGM 0.017 lg/kg bw/day

91939491

8888

Bushnik et al.(2010)

Germany 599 (3–14 years)137 (3–5 years)145 (6–8 years)149 (9–11 years)168 (12–14 years)

GM 2.66 ng/mL median 2.74 ng/mLGM 3.55 ng/mL median 3.53 ng/mLGM 2.72 ng/mL median 2.81 ng/mLGM 2.22 ng/mL median 2.13 ng/mLGM 2.42 ng/mL median 2.60 ng/mL

GM 0.060 lg/kg bw/day 9999999998

Becker et al.(2009)

Germany 147 <0.3–9.3 ng/mL Median 0.030 lg/kg bw/day Völkel et al. (2008)Belgium 193

14–16 years0.1–53.4 ng/mL (0.18–32.4 lg/g cr)GM 2.22 ng/mL (1.66 lg/g cr)

GM 0.040 lg/kg bw/day 99 Milieu enGezondheid(2010)

Italy 715 (20–74 years)111 (20–40 years)157 (41–65 years)452 (66–74 years)

GM 3.59 ng/mLGM 4.31 ng/mL median 4.4 ng/mLGM 3.95 ng/mL median 3.7 ng/mLGM 3.32 ng/mL median 3.2 ng/mL

GM 0.063 lg/kg bw/daya




Galloway et al.(2010)

Korea 516 Mean 2.74 ng/mL, median 0.64 ng/mL Mean 0.055 lg/kg bw/dayb 76 Hong et al. (2009)China 419 males

503 femalesGM 1.41 ng/mL (0.72 lg/g cr)GM 0.58 ng/mL (0.23 lg/g cr)

GM 0.032 lg/kg bw/dayc

GM 0.010 lg/kg bw/dayd5844

He et al. (2009)

China 2873–24 years

GM 3.0 ng/mL (2.75 lg/g cr)0.41–198.05 lg/g cr

GM 0.060 lg/kg bw/daya 100 Li et al. (in press)

China 116 GM 1.10 ng/mL (1.03 lg/g cr) 90 Zhang et al. (2011)Vietnam 30 GM 1.42 ng/mL (1.27 lg/g cr) 100 Zhang et al. (2011)Malaysia 29 GM 1.00 ng/mL (1.93 lg/g cr) 97 Zhang et al. (2011)India 21 GM 1.59 ng/mL (2.51 lg/g cr) 100 Zhang et al. (2011)Kuwait 32 GM 1.24 ng/mL (1.09 lg/g cr) 81 Zhang et al. (2011)Japan 36 GM 0.84 ng/mL (0.67 lg/g cr) 100 Zhang et al. (2011)Korea 32 GM 2.00 ng/mL (2.53 lg/g cr) 97 Zhang et al. (2011)All Asian countries Children

AdultsMedian 0.039 lg/kg bw/daymedian 0.037 lg/kg bw/day

Zhang et al. (2011)

US 404 pregnant women Median 1.3 ng/mL < 0.36–35.2 ng/mL Median 0.027 lg/kg bw/daye 91 Wolff et al. (2008)The Netherlands 100 pregnant women GM 1.5 ng/mL (1.7 lg/g cr), median

1.2 ng/mL (1.6 lg/g cr), range < 0.26–46 ng/mL (0.1–22.7 lg/g cr)

GM 0.024 lg/kg bw/daye

median 0.019 lg/kg bw/daye82 Ye et al. (2008)

Spain 120 pregnant women Median 2.2 ng/mL Median 0.035 lg/kg bw/daye 91 Casas et al. (2011)Mexico 60 pregnant women GM 1.95 ng/mL, 0.41 – 7.47 ng/mL GM 0.034 lg/kg bw/daya 80 Cantonwine et al.

(2010)Germany 91 samples from 47

infants (1–5 months)<0.45–17.85 ng/mL 42 Völkel et al. (2011)

US 81 (23–64 months) GM 4.8 ng/mL (6.6 lg/g cr)0.4–211 ng/mL (0.5–334 lg/g cr)

Median 0.114 lg/kg bw/day 100 Morgan et al.(2011)

Spain 30 (boys 4 years) Median 4.2 ng/mL 97 Casas et al. (2011)US 90 (girls 6–8 years) GM 2.0 ng/mL (3.0 lg/g cr)

median 1.8 ng/mL<0.3–54.3 ng/mL

GM 0.033 lg/kg bw/dayf

median 0.030 lg/kg bw/dayf94.4 Wolff et al. (2007)

US 195 samples from 35children (6–10 years)

GM 3.4 ng/mL (3.4 lg/g cr)median 3.6 ng/mL (3.5 lg/g cr)<0.36–40 ng/mL (0.2–36.3 lg/g cr)

GM 0.057 lg/kg bw/daymedian 0.060 lg/kg bw/day

95 Teitelbaum et al.(2008)

a Assuming 1.4 L urine (Lakind and Naiman, 2008) and 80 kg bw (EPA Exposure Factors Handbook 2011).b Assuming 1.4 L urine (Lakind and Naiman, 2008) and 70 kg bw (Hong et al., 2009).c Assuming 1.6 L urine (Lakind and Naiman, 2008) and 70 kg bw (Hong et al., 2009).d Assuming 1.6 L urine (Lakind and Naiman, 2008) and 70 kg bw (Hong et al., 2009).e Assuming 1.2 L urine (Lakind and Naiman, 2008) and 75 kg bw (EPA Exposure Factors Handbook 2011).f Assuming 0.6 L urine and 36 kg bw (Lakind and Naiman, 2008).


2009). However, when the population investigated is sufficientlylarge, the spot sampling approach may provide sufficient statisticalpower to categorize the average population exposure to BPA(WHO, 2010).

Assuming steady-state excretion, the daily intake of BPA corre-sponds with the excretion of BPA within 24 h (Lakind and Naiman,2008). For estimating the daily BPA intake, the urinary concentra-tions of total BPA (free and conjugated after the hydrolysis of the


conjugates) (ng/mL) are multiplied with 24 h urinary output (mL)to get the daily excretion of BPA in ng/day. Since excretion of in-gested BPA into urine is essentially complete in 24 h (Völkelet al., 2002, 2005) this was assumed to be equal to the daily intake.This estimated intake can be adjusted for body weight to obtain anexposure expressed in ng/kg bw/day (Lakind and Naiman, 2008).

5.1. Urinary BPA (ng/mL) � urinary output (mL/day)/body weight(kg) = ng BPA/kg/day

Instead of adjusting for urinary output, BPA concentrations canalso be adjusted for daily creatinine excretion. However, many fac-tors contribute to the daily variability in creatinine output such asdiurnal variation, changes in the rate of glomerular filtration, bodymass, age, gender, health status, and external factors such as diet,exercise, and drug use. Since the variation in the range of creati-nine concentration in the urine may be over 1000%, while the var-iation in daily urinary volume is up to 300% (Boeniger et al., 1993),correction for urinary output is generally preferred over creatinineexcretion (Lakind and Naiman, 2008). However, the urine volumeis also related to several factors such as liquid intake, physicalexercise, and individual health and lifestyle factors (WHO, 2010).Next to the use of generic values to describe typical urinary outputspecified for age and gender, also generic values for body weighthave to be used when individual values are not available.

Daily intake calculations based on biomonitoring data allow thecomparison of individual (or group) exposures with doses that tox-icological studies have determined to be harmful. Although thesedose calculations are performed using certain assumptions (e.g.daily urine volume or creatinine excretion, uniform metabolism),

Table 6Overview of the estimated intake of BPA through multiple exposure pathways based on a

Source Country Population Daily intake of

ChildrenTotal Food Toddlers 1088–4992 ng/

Total Food USA Children 18 months–5 years

1700–2700 ng/

Dust Eastern US Toddlers 42.2–435 ng/dapercentile)

Dust Belgium Toddlers 73–975 ng/daypercentile)

Inhalation (dust-air) USA Children(18 months–5 yeras)

7.8–14 ng/day

Dental Surgery Children (>6y) 215 ng/day

AdultsTotal Food Adults 1560–10453 ng

Canned food New-Zealand Adults 570 ng/day (avpercentile)

Canned food andbeverages

Belgium Adults 1050 ng/day (a(95th percentil

Dust Eastern USA Adults 8.44–109 ng/dapercentile)

Dust Belgium Adults 29–244 ng/daypercentile)

Thermal paper USA-Japan-Korea-Vietnam

General populationOccupationalexposed

17.4–541 ng/dapercentile)1303 – 40590 npercentile)

Paper Currencies Worldwide General populationOccupationalexposed

0.0001–1.41 ng0.0007–14.1 ng

Paper other thanthermal paper

USA General population 0.1 ng/day

Dental surgery Adults 215 ng/day

they reflect real exposures, where all possible exposure sourcesare included (Needham et al., 2007). These urinary data (Table 5)show that estimated median exposures are in the range of 0.01–0.05 lg/kg body weight (bw) per day for adults and somewhathigher (0.02–0.12 lg/kgbwper day) for children. The 95th percen-tile exposure estimates are 0.27 lg/kgbwper day for the generalpopulation and higher for infants (0.45–1.61 lg/kgbwper day)and 3- to 5-year-old children (0.78 lg/kgbwper day) (WHO, 2010).

6. Overall estimation of exposure to bisphenol-A

Based on the available data from the previous chapters, it be-comes clear that the exposure to BPA from non-food sources isgenerally lower than the exposure from food by at least one orderof magnitude for most age subgroups studied. An overview of theestimated intake through different exposure pathways based ona median and worst case intake scenario is given in Table 6 for dif-ferent studies. In a median exposure scenario, food was estimatedto contribute for more than 90% to the overall BPA-exposure for allage groups of non-occupationally exposed individuals. BPA con-centrations in food from food surveys and BPA migration from foodcontact materials were considered in this assessment. Exposurethrough dust ingestion, dental surgery and dermal absorption fromthermal paper remained below 5% in normal situations, for tod-dlers, children, and adults (Table 6). Some additional potentialsources of exposure (unpackaged food and medical devices) havebeen identified, but non-food exposure to BPA is poorlycharacterized.

A comparison between the intake assessments based on expo-sure from food and non-food source and biomonitoring values

median intake scenario.

BPA Contribution to medianexposure scenario

Refs.

day >90% Von Goetz et al.(2010)

day (median) 99% Wilson et al. (2007)

y (median – 95th <1% Loganathan andKannan (2011)

(median – 95th <5% Geens et al. (2009)

<1% Wilson et al. (2007)

<5% Von Goetz et al.(2010)

/day >90% von Goetz et al.(2010)

erage)–6900 (99th Thomson andGrounds (2005)

verage)–6050 ng/daye)

>90% Geens et al. (2010)

y (median – 95th <1% Loganathan andKannan (2011)

(median – 95th <5% Geens et al. (2009)

y (median – 95th

g/day (median – 95th

<5% Liao and Kannan(2011a)

/day (median)/day (median)

<1% Liao and Kannan(2011b)

<1% Liao and Kannan(2011a)

<5% Von Goetz et al.(2010)

0

20

40

60

80

100

120

biomonitoring canned food non-food

BPA

exp

osur

e (n

g/kg

bw

per

day

)

Fig. 3. Comparison between BPA exposure calculated from biomonitoring data andBPA exposure from canned food and non-food sources. Error bars representstandard deviations.


indicates that in general it is possible to rely on the biomonitoringdata to assess overall human exposure to BPA. Fig. 3 cumulates BPAexposure calculated from canned food/drinks, as mean/median/GMvalues taken from WHO (2010) and background documents where-in and from the specific national studies (see Chapter 2, WHO,2010). Non-food sources include exposure from dust, thermal pa-per, medical devices and dental materials (see Chapter 3, WHO,2010). BPA cumulative exposure was calculated based on biomon-itoring data (see Chapter 6, WHO, 2010).

7. Epidemiological studies

The majority of epidemiological studies used cross-sectional de-signs with a single measurement of urinary BPA. Cross-sectionalstudies have a limited interpretability, especially for outcomes thathave a long latency period (e.g. cardiovascular disease, diabetes).Moreover the use of single urine samples is another limitation gi-ven the short half-life of BPA. The association between BPA expo-sure and end-points such as cancer, reproductive outcomes,cardiovascular disease and diabetes, pubertal development out-comes and growth and neurodevelopment outcomes are summa-rized in a report of a Joint FAO/WHO expert meeting (WHO, 2010).

Three epidemiological studies revealed the association betweenhigher urinary concentrations of BPA and lower semen quality,however in two of these studies, this correlation was not signifi-cant. No evidence was found for the association between the BPAconcentrations in urine and an altered age of pubertal onset ingirls. A prospective study of Braun et al. (2009) suggested that pre-natal BPA exposures, especially during early pregnancy, may beassociated with the later development of externalizing behaviours,such as aggression and hyperactivity, and this, particularly in girls.However, replication of this study with serial measurements of uri-nary BPA is necessary.

Based on the cross-sectional analysis of urinary BPA concentra-tions from the US-NHANES, an association was reported betweenBPA-exposure and self-reported diagnosis of pre-existing cardio-vascular disease (Lang et al., 2008; Melzer et al., in press) and dia-betes (Melzer et al., 2010). Also here, confirmation withprospective studies with serial measurements of BPA is necessaryand this during the relevant windows of exposure, years or evendecades before the development of cardiovascular disease, diabe-tes and reproductive abnormalities. The question however remainsas to whether human epidemiological studies will enable to linkBPA exposure and long term effects, despite ubiquity, short half lifeand mixture effects of BPA.

8. General discussion and recommendations

The above chapters provide a detailed overview of the principalsources for human exposure to BPA and their contribution to theoverall exposure. The major human route of exposure to BPA hasbeen shown by several assessments to be the dietary pathway.However, other exposures, e.g. dermal exposure including fromthermal or recycled papers, dental materials and other medical de-vices, need to be more thoroughly characterized. Exposure throughair and dust inhalation is considered negligible. These tentativelypostulated ways of exposure have to be confirmed by biomonitor-ing data obtained from urine samples, the most suitable matrix toimprove our knowledge on absorption, distribution, metabolismand excretion of BPA.

Several characteristics of BPA deserve special attention andmake its risk assessment particularly challenging. Firstly, BPA isubiquitous because it is manufactured in large amounts and usedin a large variety of applications. Hence, BPA can be found abovethe detection limits in urine in the majority of the people moni-tored worldwide (WHO, 2010). In addition, the exposure routesare multiple. On the other hand, BPA is readily metabolized so thatthere is a continuous competition between absorption and elimi-nation within the human body. Determining a causal link betweenBPA exposure and negative health effects under such dynamic con-ditions is a major challenge for the future.

Epidemiological studies are difficult to interpret due to the fol-lowing reasons: (1) humans are exposed to numerous and variedendocrine disruptors, it is thus difficult to identify the specific ef-fects of BPA; BPA has the ability to interact with human estrogenreceptors of both a, b, and c subtypes, and in vitro experimentshave revealed significant estrogen and androgen activity of BPA.In addition, thyroid hormone receptors, PPAR-gamma receptorsand GPR30 receptors could be involved; (2) Endocrine disrupters’impact on sexual development, reproduction potency, health(especially cancers of sexual organs but also cardiovascular dis-eases and diabetes) depends upon time windows of exposure (inutero, newborn babies, adolescents, adults, menopaused women,. . .). Early exposure is more likely to account for effects due to vul-nerability of mechanisms during early set up of homeostasis ofprocesses like control of reproduction and energy balance (Bour-guigon and Parent, 2010).

Some precautions must be taken when designing biomonitoringcampaigns or protocols for epidemiological research. The followingpoints deserve particular attention for such studies:

� Attention is required to avoid external contamination with BPAduring sampling and analysis, particularly when measuring freeBPA. The nature and potential contribution of BPA sources dur-ing sampling and analysis of biological specimens is needed. Adetailed description of the sample collection protocols, includ-ing sampling location and procedures, sample handling andstorage conditions, should be included in all biomonitoringstudies. To monitor for potential external contamination, labo-ratory, as well as field blanks, are required.� The selection of several target populations throughout life in

biomonitoring studies: (i) adults, teenagers, infants, babies, pre-mature newborn babies; (ii) male or female; (iii) pregnantwomen; (iv) fertile or non-fertile men or women; (v) ethnicalgroup; (vi) identify geographical differences; (vii) different bodymass index.� Because of BPA’s short elimination half-life, strategies to

address the large variability in BPA concentrations of spot urinesamples need to be developed to adequately categorize expo-sure as appropriate to the end-point of interest. When the pop-ulation investigated is sufficiently large (e.g. nation-wide), the


spot sampling approach may provide enough statistical powerto categorize the average population exposure to BPA. For otherpurposes, biomonitoring data would be strengthened with thecollection of multiple spot urine samples, particularly in studiesaimed at evaluating the potential impact of exposure to BPA onhuman health. Furthermore, the study design should considerthe impact of time of day of sampling (e.g. in relation to con-sumption of food) and time of last urination as important expo-sure contributors to provide the best approach for BPA exposureassessment.� Focus on all exposure routes: (i) occupational exposure (plastic

industry); (ii) foodstuffs in contact with BPA containing materi-als; (iii) other oral contact materials than food (dummies, toys);(iv) dust in the (indoor) environment; (v) dermal contact (ther-mal paper); (vi) medical devices� Link between ‘‘exposure window’’ and ‘‘multiple effects’’: (i)

exposure in utero can have effects delayed until the adult periodor even on the following generations; (ii) levels of BPA in humanbody can be highly variable with time and repeated biomonitor-ing can help to control this variability; (iii) it is still unclearwhether BPA concentrations in maternal biological specimensare adequate surrogates for fetal and infant exposures.� Confounding factors and bias: (i) adjustments needed according

to age, -food habits, smoking habits, etc.; (ii) effects of otherpotential endocrine disrupting compounds and contaminants(ideally a large panel of chemicals should be monitored); (iii)unmeasured factors may confound potential BPA-outcome asso-ciations and bias effect estimates from epidemiological studies.This concern may be overstated in cases when the confoundingfactor is not associated with BPA exposure or the outcome.� To avoid some of these pitfalls, it is recommended to pay atten-

tion to the amount and quality of information that could beobtained, e.g., through a specific and detailed questionnaire ondiet, kitchen utensils and other implements, housing, occupa-tional exposure, hobbies, etc. In addition, in order to refine theexposure assessment and to improve the risk assessment, thefollowing points should be considered:� When considering all the exposure routes, more information is

needed on the bioavailability of BPA. Therefore more researchwork is needed on the absorption after dermal contact and dustinhalation.� Since a lot of contact materials are able to release BPA, it is also

important to rely on experimental procedures for an accurateand sensitive control of the quality of packaging material andkitchen utensils. It is, indeed, known that temperature, natureof the simulant and aging of the material can all affect theamount of released BPA (Nam et al., 2010). Standardized proce-dures adapted to this kind of contaminants and packagingmaterials should be developed not only for pre-market controlof kitchen implements but also for refining the prediction ofexposure.

Several additional issues highly relevant for future researchhave been identified and need to be better addressed:

� Surveys of BPA concentrations in infant formula and in toddlerfood, especially if such food is packed in metal cans� Studies on BPA migration from paper packaging to food, espe-

cially if recycled paper is used� More data on BPA concentrations in unpackaged foods have to

be gathered, together with data on the consumption patternsfor materials and products containing BPA� It would certainly be more helpful, than simply examining BPA

exposure, to try to correlate epidemiological studies with con-tamination of consumed food and drinks established usingin vitro methods (target cells equipped with reporter genes,

receptor assays) of measurements of estrogenic activity in orderto consider the whole oral exposure to endocrine disruptors.� There are several biomonitoring studies which provide scien-

tific indications of the gestational BPA exposure (Cantonwineet al., 2010; Braun et al., 2011b). It is advisable for this groupof vulnerable people (e.g. pregnant women) to employ the pre-cautionary principle and to make efforts to reduce the exposureto certain consumer products which are known to contain BPA-based polymers. It is yet unclear at the moment what the ben-efits of such reductions are. Since BPA is a potential endocrinedisruptor, the exposure of the fetus to endocrine disruptingchemicals is of high concern and should be carefully evaluatedand minimized.

In addition, the risk assessment is made very difficult becauseBPA is a potential endocrine disruptor. Therefore the risk identifi-cation and risk characterization processes are not straightforward.Indeed, numerous molecular targets can be involved and the deter-mination of a toxic threshold is not always possible. Furthermore,the toxicological tests are not yet fully validated nor worldwide ac-cepted as it is the case for in vitro and in vivo tests used to charac-terize carcinogenic/mutagenic compounds. This critical point hasto be solved in order to convince public authorities and industrialcompanies of the risks of endocrine disruptors for the publichealth, the future of the human population and the environment.Therefore, as in the case of carcinogenic agents, epidemiologicalstudies on a large scale may be necessary to obtain more evidencefor deleterious effects.

Another area of concern in the effects of EDCs in general and ofBPA in particular is the possible transgenerational mode of actiondue to alterations of the epigenome during exposure in early life(Bernal and Jirtle, 2010). Such mechanisms are challenging the epi-demiological studies and reinforce the question as to whether pre-venting measures should be proposed in pregnant women andyoung children, following a precautionary principle.

Finally, it is important to be aware that BPA is not the onlychemical of concern. There are many alternatives to BPA for whichthe toxicological properties are not known. Therefore, it could beadvisable to recommend the use of multi-contaminant analysismethods, on the one hand, so that a large panel of potential endo-crine disrupting compounds could be monitored in the same time,and also the use of biological screening methods, on the otherhand, in order to be able to detect the presence of still unknownchemicals with endocrine disrupting potential.

Conflict of Interest

The authors declare that there are no conflicts of interest.

Acknowledgments

AC and TG acknowledge Funds for Scientific Research (FWO)and University of Antwerp for financial support. The authors thankthe Belgian Superior Health Council for initiating the study and forthe financial contribution.

References

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