Analytical and environmental aspects of the flame retardant tetrabromobisphenol-A and its derivatives

Journal of Chromatography A, 1216 (2009) 346–363

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Journal of Chromatography A journal homepage: www.elsevier.com/locate/chroma

Review

Analytical and environmental aspects of the flame retardant tetrabromobisphenol-A and its derivatives Adrian Covaci a,∗ , Stefan Voorspoels b , Mohamed Abou-Elwafa Abdallah c , Tinne Geens a , Stuart Harrad c , Robin J. Law d a

Toxicological Centre, Department of Pharmaceutical Sciences, University of Antwerp, Universiteitsplein 1, B-2610 Wilrijk, Belgium Institute for Reference Materials and Measurements (IRMM), European Commission, Joint Research Centre, Retieseweg 111, B-2440 Geel, Belgium Division of Environmental Health and Risk Management, School of Geography, Earth and Environmental Sciences, University of Birmingham, Birmingham B15 2TT, UK d The Centre for Environment, Fisheries and Aquaculture Science, Cefas Burnham Laboratory, Remembrance Avenue, Burnham on Crouch, Essex CM0 8HA, UK b c

a r t i c l e

i n f o

Article history: Available online 14 August 2008 Keywords: Tetrabromobisphenol-A TBBPA Analytical methods Environmental levels Human exposure Regulatory aspects Review

a b s t r a c t The present article reviews the available literature on the analytical and environmental aspects of tetrabromobisphenol-A (TBBP-A), a currently intensively used brominated flame retardant (BFR). Analytical methods, including sample preparation, chromatographic separation, detection techniques, and quality control are discussed. An important recent development in the analysis of TBBP-A is the growing tendency for liquid chromatographic techniques. At the detection stage, mass-spectrometry is a well-established and reliable technology in the identification and quantification of TBBP-A. Although interlaboratory exercises for BFRs have grown in popularity in the last 10 years, only a few participating laboratories report concentrations for TBBP-A. Environmental levels of TBBP-A in abiotic and biotic matrices are low, probably due to the major use of TBBP-A as reactive FR. As a consequence, the expected human exposure is low. This is in agreement with the EU risk assessment that concluded that there is no risk for humans concerning TBBP-A exposure. Much less analytical and environmental information exists for the various groups of TBBP-A derivatives which are largely used as additive flame retardants. © 2008 Elsevier B.V. All rights reserved.

Contents 1.

2.

TBBP-A—general information . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.1. Production and usage volumes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.2. Applications . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.3. Regulatory aspects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.4. Toxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.5. TBBP-A derivatives . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.6. Strategy of the review . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Analytical methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.1. Physico-chemical properties of TBBP-A . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2. Extraction and clean-up . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2.1. Abiotic samples . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2.2. Biological samples . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2.3. Fractionation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3. GC–MS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.4. LC–MS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.5. Capillary electrophoresis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.6. Quality assurance . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.7. Analytical methods for TBBP-A derivatives . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

∗ Corresponding author. Tel.: +32 3 820 2704; fax: +32 3 820 2722. E-mail address: [email protected] (A. Covaci). 0021-9673/$ – see front matter © 2008 Elsevier B.V. All rights reserved. doi:10.1016/j.chroma.2008.08.035

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Environmental levels (TBBP-A and derivatives) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1. Abiotic matrices . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1.1. Air . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1.2. Indoor dust . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1.3. Water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1.4. Soil . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1.5. Sewage sludge and sediment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2. Biological matrices . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.3. Food . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.4. Humans . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Concluding remarks . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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1. TBBP-A—general information 1.1. Production and usage volumes Tetrabromobisphenol-A (TBBP-A) is currently produced in the USA, Israel and Japan, but not in the EU [1,2]. The industrial production process involves the bromination of bisphenol-A with bromine in the presence of a solvent, such as methanol or a halocarbon, 50% hydrobromic acid or aqueous alkyl monoethers. Due to the nature of the process and the by-products (hydrobromic acid and methyl bromide) that can be formed, the production process is largely conducted in closed systems [3]. TBBP-A was reported as the brominated flame retardant (BFR) with the highest production volume, covering around 60% of the total BFR market [4,5]. Yet the most recent information released by the Bromine Science and Environmental Forum (BSEF) is based upon the market demand figures for 2001 [5]. The global consumption estimates of TBBP-A vary from 120,000 [6] to 150,000 tons/year, including TBBP-A derivatives [7]. At that time (2001), Asia registered the highest consumption of TBBPA (89,400 tons/year) followed by the Americas (18,000 tons/year) and Europe (11,600 tons/year). The European BFR Industry Panel (EBFRIP) reported the size of the global TBBP-A market to be 170,000 tons in 2004 [8]. They also state that the market is increasing (Fig. 1) and that a shift in the consumption volume can be observed towards Asia. TBBP-A can be imported into a country in various forms, either as a primary product or in finished or partially finished products. Examples include plastics, printed circuit boards and electronic equipment. These imports may be an important source of TBBPA in the EU, but limited information is available. The TBBP-A risk assessment estimated the imported amount of TBBP-A in the EU as primary product to be 13,800 tons/year, in partially finished products (e.g. polymers, epoxy resins) around 6000 tons/year and in

Fig. 1. TBBP-A market size in metric tons during 1995–2004 (as given in Ref. [8]).

finished products around 20,200 tons/year [1,2]. The total amount of TBBP-A imported into the EU was estimated to be around 40,000 tons/year [1,2]. 1.2. Applications BSEF [5] reported that 58% of TBBP-A is used as a reactive FR in epoxy, polycarbonate and phenolic resins in printed circuit boards, 18% is used for the production of TBBP-A derivatives and oligomers, while 18% is used as additive FR in the manufacture of acrylonitrile–butadiene–styrene (ABS) resins or high impact polystyrene (HIPS). However, BFR industry spokespersons claim that, since it was not effective, TBBP-A was never used as an additive FR in HIPS [1,2], while the European Flame Retardants Association indicates that TBBP-A is “possibly” used in HIPS [9]. TBBP-A is used primarily as an intermediate in the manufacture of epoxy and polycarbonate resins, where it becomes bound covalently in the polymer and is thus an integral part of the product. The only potential for exposure that remains originates from un-reacted TBBP-A, if an excess has been added during the production process. TBBP-A is also used as a reactive FR in polycarbonate and unsaturated polyester resins. Polycarbonates are used in communication and electronics equipment, electronic appliances, transportation devices, sports and recreation equipment, lighting fixtures and signs. Unsaturated polyesters are used for making simulated marble floor tiles, bowling balls, furniture parts, coupling compounds for sewer pipes, automotive patching compounds, buttons, and for encapsulating electrical devices. Commercial FR epoxy resins contain up to approximately 20% bromine. The main use of these resins is in the manufacturing of rigid epoxy laminated printed circuit boards. There are two main types of rigid or reinforced laminated printed circuit boards that are commonly used [10]. These are usually either based on glass fiber reinforced epoxy resin (designated FR4) or cellulose paper reinforced phenolic resin (designated FR2), but a range of type is available. The FR4-type laminate is by far the most commonly used laminate and is typically made by reaction of around 15–17% TBBP-A in the epoxy resin [10]. The most commonly used laminate is approximately 1.6-mm thick and the TBBP-A content has been estimated at around 0.42 kg/m2 [10]. This type of laminate is used typically in computers and telecommunications equipment. As an additive FR, TBBP-A is generally used with antimony oxide for optimum performance [6]. Antimony oxide is not used generally in conjunction with TBBP-A in reactive FR applications [1,2]. TBBPA is considered an alternative additive FR to octabromodiphenyl ether (OctaBDE) mixture in ABS. The use of OctaBDE in this application is no longer allowed in the EU [11]. It is therefore possible that the amount of TBBP-A used in this application in particular could increase in the future. As additive FR, it does not react chemically with the other components of the polymer, and there-

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fore may leach out of the polymer matrix after incorporation, with important implications for human exposure. Concentrations of TBBP-A commonly found in these applications are between 10% and 20% (by weight), depending on the polymer. ABS resins are used in automotive parts, pipes and fittings, refrigerators, business machines, and telephones. HIPS resins are used in packaging, consumer products, electrical and electronic equipment, furniture, building and construction materials [3]. The largest additive use of TBBP-A is found in television casings [1,2]. Other uses include: PC monitor casings, components in printers, fax machines and photocopiers, vacuum cleaners, coffee machines and plugs/sockets. TBBP-A is also used in the manufacture of derivatives, some of which are also used as flame retardants, generally in niche applications, see Section 1.5 [3]. A large number of organobromine compounds, such as bromophenols, are naturally produced in the environment, many by marine organisms [12]. In particular bis(3,5-dibromo4-hydroxyphenyl)methane, structurally similar to TBBP-A, is produced by the segmented marine worm Thelepus setosus [12]. However, TBBP-A itself has not yet been identified as being produced by natural sources. 1.3. Regulatory aspects Currently, there are no restrictions on the production of TBBPA or its derivatives. In 2003, an EU Directive on the handling of waste electrical and electronic equipment (WEEE) [13] was adopted which contains the following elements: (i) the EU Member States shall setup separate collection schemes and ensure the proper treatment, recovery and disposal of WEEE; (ii) the treatment, recovery and disposal of WEEE shall be financed by producers to create economic incentives to adapt the design of electrical and electronic equipment to the prerequisites of sound waste management; (iii) Consumers shall have the possibility to return their equipment free of charge and need to be informed about the possibilities of returning WEEE; (iv) The WEEE Directive requires selective treatment of plastics containing BFRs, including TBBP-A. In Europe, TPPB-A is on the fourth list of priority chemicals [14] foreseen under European Council (EC) Regulation No. 793/93 of 23 March 1993 regarding the evaluation and control of the risks of existing substances. REACH is a recently implemented EU regulation on chemicals and their safe use, which deals with the registration, evaluation, authorization and restriction of chemical substances and entered into force on 1st June 2007 [15]. In the context of the REACH legislation, TBBP-A will be one of the first substances to go through the registration procedure due to its high production volume [5]. All the necessary studies for REACH registration are already developed in the context of the EU risk assessment. The EU risk assessment of TBBP-A on human health (Part II) concluded that there was no human health hazards of concern and no risks were identified [2]. This preliminary report identified no risk for TBBP-A when used as a reactive BFR, such as in the epoxy resins of printed circuit boards. However, the EU environmental risk assessment for TBBP-A confirmed a risk in some scenarios for surface water, sediment and soil when TBBP-A is used as an additive in ABS plastics [1]. BSEF indicates that risks from additive application are manageable through a Voluntary Emissions Control Action Program (VECAP), to which 89% of European customers who use TBBP-A in an additive application have agreed to reduce their emissions [8]. Many uncertainties remain regarding the risk assessment, in particular concerning the emission estimates and the biodegradation rates in the environment [16]. It should be noted that the risk assessment report also identified a risk if sludge containing TBBP-A is applied to agricultural land (see next paragraph). Industry

believes, however, that this does not happen and that sludge from user sites is either sent for incineration or to controlled landfills [5]. The EU environmental risk assessment also concluded that there is a need for further information and/or testing [1]. It is possible that TBBP-A degrades to bisphenol-A during anaerobic sewage sludge treatment processes (which could lead to bisphenol-A being applied to soil), or in anaerobic freshwater and marine sediments [1,17]. The potential risks to sediment and soil have been assessed in the updated risk assessment of bisphenol-A, for both reactive and additive flame retardant uses [17]. Currently, TBBP-A is not listed in the EU Water Framework Directive, which came into force in January 2007 [18]. Hence, there are no European monitoring schemes running currently to assess the presence of this chemical in the European water bodies. Another possible metabolite/degradation product – TBBP-A dimethyl ether – may meet the screening criteria for a persistent, bioaccumulative and toxic (PBT) substance, albeit using mainly estimated data. Its presence has been investigated in some recent studies of anaerobic transformation in freshwater aquatic sediment and sewage sludge, and anaerobic and aerobic soil transformation [1]. Although inconclusive, the results suggest that it is a very minor degradation product. Given that a need for risk reduction measures has already been identified for some uses (which should reduce the environmental burden of the parent compound), no further specific work is recommended to address this issue at the present time. The risk characterization for the marine environment indicates a possible risk from some applications. A similar reduction mandate was enacted in Japan in 2001 (Recycling of Specified Kinds of Home Appliances), and one is currently being established in China for electronic and electrical waste (Regulations on Recycling and Disposal of Waste and Used Household Electrical Appliances) [19]. The Ministry of Japan included TBBPA in their environmental surveillance program beginning in 2003 [20]. China is currently preparing legislation on WEEE, similar to the EU Directive concerning WEEE 2002/96/EC, which is timely as China is becoming a major recipient of electronic and electrical waste [21]. In North America the legislative focus is still firmly on the polybrominated diphenyl ethers (PBDEs), while TBBP-A has received little attention to date. Canada is in the process of assessing the human and environmental risks of TBBP-A and its diglycidyl and allyl ether derivatives [22]. In June 2005, the Australian Ministry for Health and Ageing declared TBBP-A as a “Priority Existing Chemical”. In the near future, TBBP-A will therefore be subjected to an assessment of its potential effects to human health and the environment, which will be conducted under the National Industrial Chemicals Notification and Assessment Scheme. The Australian authorities are currently compiling from importers and producers information on quantities and use of these substances as well as general scientific information which is already available. No regulatory actions are under consideration elsewhere [5]. 1.4. Toxicity Oral administration studies with rats and mice indicate that TBBP-A has a low acute toxicity, with LD50 >5 and >4 g/kg for rats and mice, respectively [3]. Due to the structural resemblance to the thyroid hormone thyroxin (T4 ) and bisphenol A, a suspected endocrine disruptor [17], the major concern regarding TBBP-A is its potential as an endocrine disruptor. The thyroid hormonal activity of TBBP-A was examined by Kitamura et al. [23] using rat pituitary GH3 cell lines, in which release of growth hormone is thyroid hormone-dependent. In those experiments, TBBP-A stimulated the production of growth hormone and enhanced the prolifera-

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Fig. 2. Chemical structures of TBBP-A (a) and principal derivatives: TBBP-A dimethyl ether (b), TBBP-A bis(2,3-dibromopropyl ether) (c), TBBP-A bis(allyl ether) (d), TBBP-A bis(2-hydroxyethyl ether) (e), TBBP-A brominated epoxy oligomer end-capped with epoxy groups (f), TBBP-A brominated epoxy oligomer end-capped with tribromophenol (g), TBBP-A carbonate oligomer end-capped with phenyl groups (h) and TBBP-A carbonate oligomer end-capped with tribromophenol (i).

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Table 1 Main uses of TBBP-A derivative flame retardants Compound

Use

Tetrabromobisphenol-A dimethyl ether Tetrabromobisphenol-A dibromopropyl ether

Not produced commercially Additive flame retardant in polyolefins and copolymers such as polyethylene, polypropylene and polybutylenes Reactive flame retardant in polystyrene foams An additive flame retardant in engineering polymers, epoxy resins, polyesters, polyurethane, laminates for electronic circuit boards and adhesives and coatings Reactive flame retardant in high-impact polystyrene, ABS, ABS/polycarbonate, polybutylene terephthalate-alloys, polybutylene terephthalate and thermosetting resins Additive flame retardant in ABS and engineering thermoplastics

Tetrabromobisphenol-A bis(allyl ether) Tetrabromobisphenol-A bis(2-hydroxyethyl ether) Tetrabromobisphenol-A brominated epoxy oligomer Tetrabromobisphenol-A carbonate oligomer

tion of GH3 cells. TBBP-A similarly also enhanced proliferation of the rat pituitary MtT/E-2 cell lines, whose growth is estrogendependent. These results suggest that TBBP-A acts both as a thyroid hormone and estrogen agonist [23]. These findings are similar to those of Ghisari [24] who observed a growth of GH3 cells which could not be counteracted by the inhibiting growth effect of the anti-estrogen fulvestrant. These data also indicate that the effect of TBBP-A is thyroid hormone-like and estrogen receptormediated [24]. In an in vitro study performed by Kester et al. [25], TBBP-A proved to be a rather potent inhibitor of the sulfation of estradiol by estrogen sulfotransferase, an important inactivation pathway of estradiol. Inhibition of this enzyme may lead to increased bioavailability of estradiol in vivo. The resulting weak estrogen-like properties have been confirmed by several other studies [26–28]. While TBBP-A produced a thyreomimetic effect on the GH3 pituitary cell line (see above), an anti-thyroidal effect was observed on Chinese hamster ovary cells transiently transfected with T3 receptors, as well as an inhibition of the binding of triiodothyronine (T3 ) to thyroid hormone receptors [29]. Moreover, TBBP-A was shown to be a potent in vitro inhibitor for the binding of T4 to transthyretin, the thyroid hormone-binding transport protein in plasma. The binding of TBBP-A is 10 times stronger than that of the natural ligand T4 [30,31]. TBBP-A is also immunotoxic, as demonstrated by in vitro inhibition of the expression of CD25, a receptor essential for the proliferation of activated T-cells [32]. Further, TBBP-A neurotoxicity was determined by inhibition in vitro of neurotransmitter uptake into synaptosomes and dopamine uptake into synaptic vesicles [33] and generation of free radicals [34]. Additionally, TBBP-A has been shown to interfere with cellular signaling pathways [35]. Toxicological effects, such as severe disorientation, lethargy, decreased egg production and decreased reproductive success were observed in a partial life-cycle test with zebrafish (Danio rerio) exposed to environmentally relevant water-borne concentrations of TBBP-A [36].

1.5. TBBP-A derivatives TBBP-A is also used in the manufacture of derivatives [1,2]. The main derivatives (Fig. 2) produced from TBBP-A are TBBP-A dibromopropyl ether, TBBP-A bis(allyl ether), TBBP-A bis(2-hydroxyethyl ether), TBBP-A brominated epoxy oligomer, and TBBP-A carbonate oligomers [3]. The main use of these derivatives is as flame retardants, usually in specialized (or niche) applications (Table 1). These derivatives may be used as either reactive or additive intermediates in polymer manufacture. TBBP-A bis(2,3-dibromopropyl ether) is used as an additive FR in polyolefins and copolymers, such as high density polyethylene, low density polyethylene, polypropylene and polybutylenes [3,37,38]. In general, product loadings vary between 1 and 10% (by weight) in polypropylene [39]. Loadings can be reduced when used in con-

junction with antimony oxide. Flame retarded polypropylene is used in building applications (mainly in pipes for water discharge, but also film and sheet for roofing), textiles, and in electrical and electronic applications such as wire nuts, lamp sockets, coil bobbins, connectors, wire and cable, housings of electrical appliances or TV yokes [40]. De Schryver et al. [41] reported that it could also be used in high impact polystyrene at 5% by weight. TBBP-A bis(allyl ether) is used as a reactive FR in polystyrene foams (expandable polystyrene-EPS) [3,37,38]. There is a major lack of data for this derivative. TBBP-A bis(2-hydroxyethyl ether) is used as an additive FR in engineering polymers (e.g. polybutylene terephthalate and polycarbonate), epoxy resins, thermoplastic polyesters, polyurethane, laminates for electronic circuit boards, adhesives and coatings [3,37,38]. Also for this derivative, there is a major lack of data. TBBP-A brominated epoxy oligomers are also known as TBBPA diglycidyl ethers. There are two chemically different types of brominated epoxy oligomers. One has two epoxy groups at the end of the molecule, similar to epoxy resins used for printed circuit boards (EP-type). The other type, which is TBBP-A epoxy endcapped with tribromophenol (EC-type), has no reactive groups. Both types of oligomer are reactive FRs used in housings for business machinery and electrical/electronics parts based upon HIPS, ABS, ABS/polycarbonate, polybutylene terephthalate-alloys, polybutylene terephthalate and thermosetting resins. The concentrations of these FRs in ABS are around 20% (by weight). As well as fully tribromophenol end-capped oligomers, some products are available with around 50% end-capping with tribromophenol [42]. The molecular weights of the products vary between 700 and 50,000 g/mole, and differ depending on the application. TBBP-A carbonate oligomers are produced by reaction of TBBP-A with phosgene [37]. In this respect, they can be considered similar to the reactive use of TBBP-A in polycarbonates described above. These oligomers are used as additive FRs in ABS and engineering thermoplastics such as poly(butyleneterephthalate), polycarbonate, poly(ethyleneterephthalate) and phenol–formaldehyde resins [3,37]. Both phenoxy-terminated TBBP-A carbonate oligomers and tribromophenoxy-terminated TBBP-A carbonate oligomers are produced [3]. A TBBP-A diglycidyl ether-carbonate oligomer has also been reported. TBBP-A dimethyl ether (diMe-TBBP-A) is not known to be produced commercially and it is not certain whether it is used as a flame retardant [3], but it has been found in the environment [43–45]. The occurrence in the environment can be explained by the O-methylation of TBBP-A in certain biological processes [46]. Others: bis(hydroxyethyl) TBBP-A ethylene glycol can also be used to produce polyester fibers [37]. Ash and Ash [38] indicated that TBBP-A diacrylate could be used in automotive coatings and wire and cable coatings. A TBBP-A bis(2-ethylether acrylate) derivative has also been reported. The commercial significance of these products is unclear.

A. Covaci et al. / J. Chromatogr. A 1216 (2009) 346–363

1.6. Strategy of the review All available literature on TBBP-A (analytical methods, levels in the environment and humans), published until February 2008 in peer-reviewed scientific journals, conference proceedings or official reports found on the internet, was acquired and classified. Since this was not the primary scope of the review, information about the toxicology of TBBP-A was only briefly mentioned. Although several reviews are available for BFRs [47–49], specific information on TBBP-A is often “lost” in the greater dataset available for PBDEs or hexabromocyclododecanes (HBCDs). Therefore, we identified a stringent need for a comprehensive review on TBBP-A in which the available literature is critically discussed and recommendations for further research are given. Similar reviews are already available for PBDEs [50] and HBCDs [51]. 2. Analytical methods Although the literature is abundant in analytical methods for PBDEs and HBCDs, there are far fewer methods described for the determination of TBBP-A. In fact, it seems that TBBP-A has been seen merely as a potential additional analyte and not as a target compound itself, other than in a few cases. Determinations are usually accomplished using LC–MS techniques, where TBBP-A is measured together with HBCD or other phenolic (halogenated) organic compounds. Table 2 summarizes relevant data for selected analytical procedures used for the determination of TBBP-A in a wide variety of abiotic and biotic samples. Further information on various analytical methods for BFRs, including TBBP-A, has recently been reviewed [47,48]. 2.1. Physico-chemical properties of TBBP-A Due to its distinct physico-chemical properties, the determination of TBBP-A requires specific analytical approaches and these are highlighted below. The pKa1 and pKa2 values of TBBP-A are estimated at 7.5 and 8.5, respectively [3], which means that in neutral environments, a substantial part of the TBBPA is present in its dissociated state. This causes losses in the clean-up steps when a neutral environment combined with polar solvent is maintained (the polar solvent could just be a small amount of co-extracted water from the sample). Care should be taken to avoid these losses and a possible solution is to treat the raw extract with acidified water. This results in non-dissociated TBBP-A only, which is driven almost quantitatively towards the organic phase. Such behaviour is similar to that of other phenolic organohalogenated compounds (e.g. pentachlorophenol). These properties have a significant effect on the partitioning of TBBP-A in the environment and biota. Contrarily to PBDEs and HBCDs, which are neutral compounds, TBBP-A is not accumulating in fatty tissues, but it is primarily retained in blood through protein binding. 2.2. Extraction and clean-up 2.2.1. Abiotic samples Abiotic matrices reviewed include (i) water, (ii) air, (iii) soil, sediment and sewage sludge and (iv) polymers (Table 2). (i) Water. Because of the low concentrations expected in water, large volumes (up to 1000 mL) are typically required to ensure positive detection of TBBP-A [52,71]. Suzuki and Hasegawa [52] used solid-phase extraction (SPE) on Abselut Nexus cartridges as a fast and valuable technique allowing the simultaneous determination of TBBP-A (recovery 103 ± 16%) and other BFRs,

351

together with a significant reduction in the organic solvent consumption from 50 mL dichloromethane (DCM) to 5 mL acetone. A restricted access media-molecularly imprinted polymer (RAM-MIP) was used for the selective on-line pre-treatment and enrichment of TBBP-A in a river water sample, followed by separation and determination by LC–MS [54]. The betweenday precision for the assay of TBBP-A at 25 pg/mL was 1.6%. Solid-phase micro-extraction (SPME) has been investigated for the determination of brominated phenols and TBBP-A in aqueous samples [72]. The extraction procedure involves an in situ acetylation followed by SPME extraction. The studied factors were the type of fiber, extraction mode (direct immersion or headspace), exposure of the fiber directly into the sample or into the headspace over the sample and extraction temperature. The polydimethylsiloxane fiber was found the most suitable for the extraction of TBBP-A from water and the highest response was achieved in headspace mode at 100 ◦ C. The obtained limits of detection (LOD) were at the low pg/mL level. (ii) Air. For air samples, filters (glass fiber and PUFs) were extracted by sonication and analyzed without further clean-up [55]. Similarly, Inoue et al. [56] have eluted the filters used for air sampling with methanol (MeOH) and analyzed the extracts without further clean-up. (iii) Soil, sediment and sewage sludge. The sample preparation for BFR analysis, including TBBP-A, in sewage sludge and soil has been reviewed by Eljarrat and Barcelo [49] and the most common methods are given in Table 2. Soxhlet extraction, a robust, efficient and low-cost technique, is a primary option for the determination of BFRs in soils and sediments. In general, mixtures of acetone and n-hexane in different proportions (1:1 or 1:3, v/v) have been found to provide the best recoveries for TBBP-A [57]. Pressurized liquid extraction (PLE) has also been evaluated for the analysis of TBBP-A in dried soil, sediments and sewage sludge. Extraction with DCM at 100 ◦ C has been found to provide almost quantitative recoveries (∼80%) of TBBP-A [60], providing that several PLE cycles (e.g. 2 × 5 min cycle) instead of a single longer PLE cycle were carried out. An alternative procedure based on matrix solid-phase dispersion (MSPD) for sample preparation in the analysis of halogenated bisphenol derivatives in river and marine sediment and urban sewage sludge has been developed by Blanco et al. [73]. Approximately 200 mg sewage sludge or sediment were acidified with HCl, dried with Na2 SO4 , followed by mixing with 1 g C18 -modified silica and 2 g Florisil. Acetonitrile (7 mL) delivered clean extracts combined with the highest recoveries for TBBP-A (63%). A liquid–liquid extraction (LLE) followed by SPE has been developed for the simultaneous determination of halogenated bisphenol-A derivatives, including TBBP-A, in sediment and sludge samples [58]. Samples were extracted with methyl tert-butyl ether and the analytes were partitioned using an aqueous solution of NaOH. The extract was subsequently acidified, and enrichment and desalting were performed by passing the extract over a C18 cartridge and a silica filled cartridge, respectively. After clean-up, the target compounds were determined by LC–MS/MS. The method LOD for sediment and sludge for TBBP-A was 0.05 ng/g dry weight (dw). Mean recovery of TBBP-A from spiked samples was 102 ± 5%. A stir bar coated with poly(dimethysiloxane)-␤-cyclodextrin on one side has been prepared for the first time by a sol–gel method and was coupled with ultrasonic assisted extraction for the determination of TBBP-A in soil and dust samples by HPLC with diode array detection (DAD) [74]. Extraction time, desorption solvent, concentration of MeOH/acetone and NaCl in the matrix, pH, temperature and stirring speed were opti-

352

Table 2 Overview of typical analytical procedures used for the determination of TBBP-A in selected matrices Reference

0.0002

[52]

–

0.02

[53]

>95%

1–4

0.01

[54]

LC–ESI-MS/MS

75–93

–

–

[55]

–

LC–ESI-MS

87–99.5

–

0.1

[56]

–

LC–APCI-MS/MS

101

4

0.0002

[52]

LLE with H2 SO4 + GPC + SiO2

LC–ESI-MS

–

–

0.5

[57]

SPE-C18 cartridges + SPE-Si cartridges (500 mg, 3 mL) SPE Supelclean ENVI-18 cartridges (3 ml) + filtration.

LC–ESI-MS/MS

102.2

5.1

0.05

[58]

LC–ITD/MS

94

15

0.5

[59]

GC-HRMS

>80

–

–

[60]

UPLC–ESI-MS/MS

30.8–92.5

–

–

[61]

GPC + LLE with H2 SO4

LC–ESI-MS

–

–

–

[62]

LLE with H2 SO4 + GPC + SiO2

LC–ESI-MS

–

–

0.5

[57]

GPC + SiO2 –H2 SO4 + deact SiO2 (1.5% H2 O) LLE with H2 SO4

LC–ESI-MS

80–127

9–27

1–200

[63]

LC–ESI-MS/MS

79–93

6–7

0.1

[64]

LC–TOF-MS

56–94

8

20

[65]

14 –

10 1

Pre-treatment

Extraction procedure

Clean-up

Instrumental analysis

Recovery (%)

Water Landfill leachate (1000 mL)

Filtration

Abselut Nexus SPE (AcN, 5 mL) SPE-C18 cartridges (Bond Elut, 500 mg; Varian) –

–

LC–APCI-MS/MS

103

–

LC–ESI-MS/MS

>85%

–

LC–ESI-MS

Sonication (2 × 5 mL, 2 × 20 min, AcN)

Filtration

Elution with 30 mL MeOH

Sonication (AcN, 10 min) Soxhlet with Acet:Hex (3:1, 6 h or 1:1, 12 h) Soxhlet (MTBE, 12 h)

Wastewater (50 mL)

River water (2 mL) Air Air samples (3 m3 )

Air samples (10 m3 )

Soil, sediment and sewage Marine sediment (1 g)

1 mL of ascorbic acid (0.1M)—to avoid chlorination during storage Filtration, isotope imprinting

25 mm glass fiber filter and 2 PUF plugs pre-cleaned with MeOH, Acet and DCM 47 mm glass fiber filter + Empore disk (SDB-XD 47 mm/0.5 mm) Air dry

Sediments and sewage sludge

Mixed with Na2 SO4

Sediments and sewage sludge (10 g) Sediments and sewage sludge (0.1–1.8 g)

Mixed with Na2 SO4

Sediment (10 g) sewage sludge (1 g)

Freeze dry + homogenization

Soil (30 g)

–

Soxhlet (Acet:DCM 4:1, 22 h)

Mixed with Na2 SO4

Soxhlet (Acet:Hex 1:1, 4 h) Soxhlet with Acet:Hex (3:1, 6 h) or homog. by Ultra Turrax Soxhlet (Acet:Hex, 1:1, 4 h) Soxhlet (Acet:Hex 1:1, 7 h) Column extraction (Acet:cyclohexane, 1:3, 1 h)

Biological samples Harbour porpoises

–

Marine mammals, fish and marine invertebrates

Mixed with Na2 SO4

Cod muscle (5 g)

Na2 SO4

Fish (2–10 g)

Mixed with Na2 SO4

Egg (10 g)

Homogenization + Na2 SO4 (overnight)

Sonication (DCM:MeOH 1:9, 1 h) + agitation (3 h) PLE (DCM, 100 ◦ C, 12.7 MPa)

Derivatization (CH2 N2 ) + SiO2 + deact SiO2 + SiO2 –H2 SO4 + SiO2 –AgNO3 SPE C18 -cartridges (1.5 mL, 100mg) + filtration

GPC + deact Florisil (0.5% H2 O) + derivatization

GC–LRMS GC–HRMS Blood serum (10 mL)

–

Adipose tissue (0.5 g)

–

Human milk (1 g)

Freeze dry, homogenization

Human serum (5 mL)

Formic acid

Tissues of humans, dolphins and sharks (3, 20 g)

Mixed with Na2 SO4

-LLE (EtOAc, 12 + 8 mL) + LLE (AcN, 3 mL + Hex, 3 × 3 mL) -LLE (AcN, 3mL + Hex, 3 × 3 mL)

Hex layer (PBDEs): Oasis HLB SPE + SiO2 + SiO2 –H2 SO4

GC–HRMS (after sylilation)

40

RSD (%)

13

4–7

0.2–4

[66]

0.004 (wet weight) 0.0003 (wet weight)

[67]

AcN layer (HBCD + TBBP-A): enzymatic hydrolysis (50 ◦ C, 4 h) + Oasis HLB + silica SPE

SLE (Acet:DCM, 1:1, 12 + 6 mL) + LLE (AcN, 3 mL + Hex, 3 × 3 mL) Abselut Nexus SPE

SiO2

LC–ESI-MS/MS

83–104

<8

Soxhlet DCM:Hex (3:1; 16 h)

GPC + LLE with H2 SO4 + filtration

LC–ESI-MS/MS

93.2

6.3

[68]

A. Covaci et al. / J. Chromatogr. A 1216 (2009) 346–363

LOD (ng/mL, ng/g, ng/m3 )

Sample type

Acet, acetone; AcN, acetonitrile; Hex, n-hexane; DCM, dichloromethane; EtOAc, ethyl acetate; MTBE, methyl-tert-butyl ether. Solvent mixtures: proportions as v/v. PUF, polyurethane foam. Other acronyms, as identified in the body text.

[70] 790,000 – – LC-UV Ultrasonication (iso-propanol) Pulverized in mills WEEE polymer fractions

Filtration

[70] 10–100/50–100,000 – 105–122 LC-UV LC-UV-APCI-MS GPC WEEE polymer fractions (980)

LC-UV-APCI-MS

PLE (iso-octane, 3 cycles) – Polymers WEEE polymer fractions (980)

THF + DCM + mixed with silica gel + dried Dissolved in THF + filtration

Dilution + filtration

90–135

–

10–100

[69]

A. Covaci et al. / J. Chromatogr. A 1216 (2009) 346–363

353

mised. The following analytical parameters were determined for the optimised method: repeatability (5–9%), intermediate precision (10%), recovery (56–71%) and detection limits (4.2 ␮g/L). (iv) Polymers. Altwaiq et al. [75] examined different procedures for the extraction of TBBP-A and its derivatives from various polymer materials. These procedures included supercritical carbon dioxide (CO2 ), modified supercritical CO2 , solvent and Soxhlet extraction. The extraction efficiency varied according to the applied methods. The results proved the high capacity of solvents such as toluene, tetrahydrofuran and acetonitrile for the extraction of several BFRs, whereas extraction using a combination of CO2 and organic solvents was more efficient than by using supercritical CO2 alone. PLE [69] and ultrasonic solvent extraction (USE) with 2-propanol [70] have also been developed for the extraction and identification of BFRs in polymers used in electrical and electronic equipment (EEE), such as TV and PC monitor housings. For polymers containing high concentrations of TBBP-A, simpler analytical techniques have been proposed. Schlummer et al. [69] used LC–UV/MS to identify and quantify TBBP-A in post-consumer plastics from e-waste. Quantification proved to be more reproducible by UV, which was probably due to the co-extraction of polymer components that accumulate in the atmospheric pressure chemical ionization (APCI) source. The authors therefore suggest using UV for the quantification and MS for identification and validation purposes. A similar approach based on a rapid method for the extraction of various BFRs followed by LC–UV was developed by Pöhlein et al. [70]. The overall runtime required for extraction and chromatographic analysis was less than 10 min. Kikuchi et al. [76] analyzed BFRs in matrix polymers by Raman spectroscopy without any sample pre-treatment. The LOD was approximately 100 ␮g/g and analysis was only 1 min. Based on the distinctive bands, the DecaBDE technical mixture and TBBP-A could be identified. Energy dispersive X-ray fluorescence analysis, LC-UV, GC–MS and infrared spectroscopy techniques have also been evaluated for the analysis of polymers after various extraction procedures [75]. A TBBP-A analogue (tetrabromobisphenol-S-bis-(2,3dibromopropyl ether)), which contains sulfur, has been identified in polypropylene used in the manufacture of TV cabinets [77]. Various chromatographic and spectroscopic techniques were used for the correct identification of this TBBP-A derivative. The feasibility of using radiofrequency glow discharge plasma spectrometry coupled with optical emission spectrometry (rf-GD-OES) as a rapid and simple tool to directly analyze polymers containing BFRs has recently been investigated [78]. The best detection limit (0.044% Br) was achieved by measuring at 827.24 nm in a He discharge. The linearity range extended up to a bromine content of 27%. 2.2.2. Biological samples Table 2 summarizes relevant methods related to TBBP-A in biological samples. The reviewed categories include (i) biological fluids (i.e. serum and plasma) and (ii) fatty foodstuffs and animal tissues. Usually, only drying and homogenization is carried out before extraction of biological samples as with abiotic samples. Except for serum and plasma, (semi)liquid samples (e.g. eggs) are usually freeze-dried and then treated as any other solid biotic sample. In general, similar extraction techniques and solvents are used for TBBP-A analysis in abiotic and fat-containing matrices, and the main differences between both sets of analytical protocols refer only to the subsequent clean-up steps.

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(i) Serum and plasma. Direct solvent shaking with ethyl acetate and acetonitrile has been demonstrated to provide relatively low, but reproducible recoveries for TBBP-A (40%), despite the high number of manipulation steps [66]. One of the main limitations of these LLE-based procedures is the long settling time or centrifugation required for phase separation. Alternative SPE-based methods [79,80] have also been proposed. These methods proved to be less laborious and allowed reduced solvent consumption and processing time, the possibility of miniaturization, and parallel sample preparation, which increases sample throughput. Hayama et al. [67] have setup a method for the determination of TBBP-A in human serum using SPE and LC coupled to electrospray interface (ESI) MS/MS (Table 2). 13 C-labelled TBBP-A was a suitable surrogate standard for the reproducible determination of TBBP-A. Method limit of quantification (LOQ) was as low as 4 pg/g serum. (ii) Biological tissues and fatty foodstuffs. Similar extraction techniques as used for abiotic samples have been used for determination of TBBP-A in complex biological tissues and fatty foodstuffs. Binary solvent mixtures typically containing acetone:n-hexane [45] or DCM:n-hexane [65] have been preferred for Soxhlet-based extractions. These techniques have a number of advantages, such as minimum sample pre-treatment required, simplicity, and high recoveries (>80%) [57]. 2.2.3. Fractionation For specific applications, isolation of the target analytes from other organohalogen compounds present in the extract can be mandatory so as to minimize interferences during the final determination step. Deactivated silica gel has been successfully applied for the quantitative separation of TBBP-A from PBDEs. In this case, iso-octane was used for the elution of PBDEs, while a more polar solvent, i.e. diethyl ether:iso-octane (15:85, v/v), was required to elute TBBP-A [57]. Florisil (activated at 450 ◦ C for 12 h and subsequently deactivated with 0.5% H2 O, w/w) has also been successfully used to separate neutral organohalogen compounds from phenolic analytes, such as TBBP-A [65]. In this case, neutral compounds were first eluted with mixtures of DCM:n-hexane (1:3, v/v), while polar mixtures of acetone:n-hexane (15:85, v/v) and methanol:DCM (12:88, v/v) were needed to elute phenolic analytes [65]. Sorbents, such as Oasis HLB® , have been reported for the fast separation of TBBP-A from HBCD diasteroisomers [66]. DCM:n-hexane (1:1, v/v) was used to elute HBCDs from the SPE cartridge, while TBBP-A was subsequently eluted with DCM [66]. 2.3. GC–MS Among the predominant BFRs, TBBP-A is the most polar molecule, and thus requires different and more complicated methods so that a proper determination can be carried out. Acidification and derivatization are compulsory before GC analysis can be carried out [80]. A GC-HRMS method requiring derivatization with methylchloroformate was developed by Berger et al. [65]. However, this method suffered from a rather restricted linear range and low recoveries due to incomplete derivatization, which was explained by the presence of bulky bromine substituents adjacent to the two hydroxyl groups. Although the GC–MS method showed better separation properties and was more sensitive for standard solutions, LC–ESI-time-of-flight (TOF)-MS was superior for the quantification of egg extracts, with a satisfactory LOQ of 3 pg TBBP-A on-column [65]. This was explained by the high-resolution filtering potential

of the TOF-MS, which minimizes the matrix background from the analyte’s mass chromatograms. Comprehensive two-dimensional gas chromatography (GC × GC) coupled with ␮ECD or TOF-MS, a technique that offers excellent separation power, has also been evaluated for the analysis of PBDEs and possible co-eluants [82]. It was found that the second dimension GC column improves the separation of Me-TBBP-A and TBBP-A from any co-eluting PBDE congeners (BDE 154 and BDE 153, respectively, when the first dimension GC is a DB-1 or DB-5 column). 2.4. LC–MS In a review of the available methods for the determination of BFRs, Covaci et al. [48] indicated that LC–MS could be applicable for the simultaneous analysis of TBBP-A with HBCDs. LC has the advantage that no derivatization step is required before determination of TBBP-A, whereas this step is necessary for its determination with GC. Derivatization has been reported to produce errors and/or analyte losses [81]. Table 3 summarizes the parameters used for various determination methods for TBBP-A. Frederiksen et al. [86] compared LC–MS/MS to GC–MS for the determination of TBBP-A in biota samples and concluded that LC–MS/MS is the method of choice, not only because derivatization is not needed, but also because of its higher sensitivity and better detection limits. Chu et al. [58] found that the efficiency of the LC separation and MS sensitivity for TBBP-A are largely affected by the mobile phase used. They reported a 30% increase in response following replacement of acetonitrile with MeOH in the mobile phase, together with a more stable detector baseline, which resulted in a lower LOQ. Further enhancement in response was observed upon the addition of 1 mM ammonium acetate to the mobile phase, which may be due to ionization enhancement. Another advantage of the LC–MS/MS determination of TBBP-A is that it enables the use of the 13 C-labelled TBBP-A as surrogate standard. This greatly enhances the quality of the analytical data obtained by compensating for any matrix-related effects that can affect analyte ion intensity [48]. A typical chromatogram obtained by LC–MS/MS for standard solutions of native and 13 C-labelled TBBP-A, together with a chromatogram for the standard reference material SRM 2585 are given in Fig. 3a and b. A single quadrupole full scan from m/z = 300 to 570 and the product ion scan for m/z = 540.8 are also given in Fig. 3c and d, respectively (Abdallah, unpublished data). Tollbäck et al. [55] reported that the most suitable LC–MS interface for TBBP-A analysis is ESI operating in negative ionization mode. ESI gave 30–40 times lower LODs compared to APCI. In addition, it permits monitoring of the intact TBBP-A molecule through the soft ionization of ESI resulting in improved method selectivity and accuracy. This finding agrees with results of Morris et al. [57]. However, for the quantification of TBBP-A in polymer fractions from WEEE using HPLC-UV/MS, Schlummer et al. [69] used the APCI source in negative ion mode, since initial trials with ESI did not produce suitable mass fragments after UV exposure. Ion-trap MS (ITD-MS) was also reported for the determination of TBBP-A in sediment and sewage sludge after LC separation [59]. The ITD scan range was set from m/z 145–543. Although ion suppression of the TBBP-A signal due to matrix components in the ESI process was not high, sewage sludge extracts suffered greatly from ion suppression and extensive clean-up was required to minimize this effect. Suzuki and Hasegawa [52] reported the analysis of HBCDs and TBBP-A in leachate samples by LC–APCI-MS. In this case, ionization by APCI yielded signal to noise (S/N) ratios two to five times higher than those obtained by ESI for HBCDs, while for TBBP-A, the S/N

Table 3 Overview of LC–MS and GC–MS parameters used for in the analysis of TBBP-A Column

Dimensions

Mobile phase (gradient)

Flow (mL/min)

Mobile phase modifiers

Ionization

Instrument

Ion

Source temp (◦ C)

Reference

LC–MS TBBP-A TBBP-A TBBP-A TBBP-A

Develosil C30 (Nomura) SunFire C18 (Waters) Cosmosil 5C18-MS-II Xterra C18 (Waters)

150 × 2 mm, 5 ␮m 150 × 2.1 mm, 3.5 ␮m 50 × 2 mm, 5 ␮m 150 × 2.1 mm, 3.5 ␮m

AcN:H2 O (y) MeOH:H2 O (y) AcN:H2 O (y) MeOH:H2 O (y)

0.2 0.2 0.2 0.2

APCI ESI ESI ESI

Triple quadrupole Triple quadrupole Single quadrupole Triple quadrupole

MRM 542.7 → 445.8 MRM 543 → 444 542 MRM 542.7 → 417.8

250 280 100 135

[52] [53] [54] [55]

Mightysil RP-18 (Kanto Chemicals) Luna C18 (Phenomenex)

150 × 4.6 mm, 5 ␮m

AcN:H2 O (y)

0.2

– – – 10 mM ammonium acetate 0.01% acetic acid

ESI

Single quadrupole

SIM 542

–

[56]

150 × 2 mm, 5 ␮m

AcN:H2 O (y)

0.25

ESI

Single quadrupole

540.9

150

[57]

Genesis C18 120A column Discovery C18 column (Supelco) and Symmetry C8 column (Waters) Hypersil GOLD C18

150 × 2.1 mm, 4 ␮m

MeOH:H2 O (y)

0.2

10 mM ammonium acetate –

ESI

Triple quadrupole

MRM 543 → 81

130

[58]

50 × 2.1 mm, 5 ␮m, 150 × 4.6 mm, 3.5 ␮m 100 × 2.1 mm, 3 ␮m

MeOH:H2 O (y) and MeOH:H2 O (n) MeOH:H2 O:AcN (y)

0.3

–

ESI

Ion-trap

Scan (145–543)

–

[59]

0.2

20 mM ammoniumacetate and 0.05 mM ammonium chloride 0.01% acetic acid 1 mM ammonium acetate

ESI

Single quadrupole

–

–

[62]

ESI ESI

Triple quadrupole TOF

MRM 541.7 → 419.8 Scan (230–550)

120 130

[64] [65]

TBBP-A TBBP-A TBBP-A TBBP-A

TBBP-A

200 × 2 mm, 3 ␮m 150 × 2.1 mm, 3 ␮m

MeOH:H2 O (y) MeOH:H2 O (y)

0.2 0.2

TBBP-A

Gemini-C18 Ace 3 C18 (Advanced Chromatography Technologies) Mightysil (Kanto)

150 × 2 mm, 3 ␮m

0.2

–

ESI

Triple quadrupole

MRM (542.7 → 445.8)

–

[67]

TBBP-A

Hypersil C18

100 × 2.1 mm, 5 ␮m

AcN:MeOH:H2 O (y) MeOH:H2 O (y)

0.25

ESI

Triple quadrupole

[68]

Nucleodur 100-C8 (Interchim) Hypersil ODS C18 (Thermo Electron)

250 × 4 mm, 5 ␮m

AcN:H2 O (y)

1

APPI

QTrap

MRM 540.9 → 79/540.9 → 81 scan

300

TBBP-A

10 mM ammonium acetate 0.1% acetic acid

n.a.

[83]

250 × 4.6 mm, 5 ␮m

MeOH:buffer (Ammonium Acetate) (n)

1

–

APCI

Triple quadrupole

Scan (150–1000)

n.a.

[69]

RP Waters Acquity BEH C18 Acquity BEH C18

50 × 2.1 mm, 1.7 ␮m

AcN:H2 O (y)

0.45

–

ESI

Triple quadrupole

MRM 542.7 → 445.8

120

[61]

150 × 2.1 mm, 1.7 ␮m

MeOH:H2 O (y)

0.5

–

ESI

Triple quadrupole

MRM 542.60 → 419.70 and 542.60 → 447.60

–

[84]

Hypersil ODS C18 (Thermo Electron)

250 × 4.6 mm, 5 ␮m (40 ◦ C)

–

–

UV (203nm)

n.a.

[69]

TBBP-A, PBDEs, DBPE

Luna 5 ␮ Phenyl-Hexyl (Phenomenex)

150 × 4.6 mm, 5 ␮m (50 ◦ C)

–

–

UV (200–400 nm)

–

–

[70]

TBBP-A-dbpe

Zorbax XDB-C18 (Agilent) SphereClone ODS 2 (Phenomenex)

150 × 4.6 mm, 5 ␮m (40 ◦ C) 250 × 4.6 mm, 5 ␮m

1 MeOH:buffer (ammonium acetate) (n) MeOH:22.5 aminoethanol:water (n) 1 AcN:H2 O (y)

–

–

DAD

–

–

[85]

–

–

UV (230, 254 nm)

TBBP-A TBBP-A

TBBP-A

UPLC–MS/MS TBBP-A TBBP-A HPLC–UV TBBP-A, PBDEs, HBCD

TBBP-A

MeOH:THF:buffer (Ammonium Acetate) (n)

1

A. Covaci et al. / J. Chromatogr. A 1216 (2009) 346–363

Compound

[69]

Compound

Column

Dimensions

Injection mode

Derivatization

Ionization

Instrument

Ion

Source temperature (◦ C)

Reference

GC–MS TBBP-A TBBP-A

DB-5 ms (J&W Scientific) UB5-P (Interchim)

30 m × 0.25 mm × 0.1 ␮m 15 m × 0.25 mm × 0.25 ␮m

Splitless Splitless

Methyl chloroformate MSTFA

EI EI

HRMS HRMS

556.7608, 554.7629 682.8509, 684.8489

275 230

[65] [66]

For acronyms, see Table 2 and text. 355

356

A. Covaci et al. / J. Chromatogr. A 1216 (2009) 346–363

Fig. 3. LC–MS/MS chromatogram of (a) a standard solution containing 20 and 50 pg on column for 13 C-labelled TBBP-A and native TBBP-A, respectively, (b) a standard reference material (SRM 2585—organic contaminants in indoor dust), (c) a single quadrupole scan from m/z = 300 to 570 and (d) the product ion scan for m/z = 540.8. Chromatographic conditions: gradient using a mobile phase of water/methanol (1:1, solvent 1) and methanol (solvent 2) at a flow rate of 150 ␮L/min, starting at 35% (solvent 2) then increased linearly to 100% (solvent 2) over 6 min, held for 5 min. Column: Varian Pursuit XRS3 C18 reversed phase analytical column (150 mm × 2 mm i.d., 3 ␮m particle size). The following multiple-reaction monitoring (MRM) transitions were as follows: 540.8-79 for the native TBBP-A and 552.8-79 for the 13 C-labelled TBBP-A.

ratio using APCI was almost half that of ESI. Recently, Debrauwer et al. [83] have investigated the applicability of LC techniques for the analysis of PBDEs. The use of atmospheric pressure photo ionization (APPI) may facilitate the analysis of PBDEs and phenolic com-

pounds, such as TBBP-A, in the same run. Without performing full optimization, LODs were found to be in the range of 200–1500 pg on column. This methodology allows to use LC–MS/MS-based methods for the identification of biotransformation products of BFRs [87].

A. Covaci et al. / J. Chromatogr. A 1216 (2009) 346–363

Recently, ultra performance liquid chromatography (UPLC)–ESIMS/MS was reported for analysis of TBBP-A in soil and food samples [61,84]. This technique combines all the advantages of LC–MS/MS in addition to shorter residence time of the analyte on-column, thus minimizing the potential for losses on-column due to adsorption and degradation. The short analysis time (4 min) can double the efficiency of the analytical method. 2.5. Capillary electrophoresis Capillary electrophoresis (CE), an efficient technique for the separation of charged species, was also found useful for the analysis of TBBP-A [73,88]. Blanco et al. [88] developed a method for the determination of TBBP-A and other phenolic compounds in environmental samples by non-aqueous CE coupled to a diodearray detector (210 nm for TBBP-A), using MeOH as solvent. Some parameters affecting the electrophoretic separation were studied, such as salt concentration, electrolyte pH, and capillary and solution temperature. Calibration curves were linear from 0.5 to 10 ng/␮L. This technique was successfully applied for the analysis of TBBP-A in water samples. Method LOQ was approximately 12 pg/␮L. 2.6. Quality assurance TBBP-A tends to adsorb to glass when using n-hexane as solvent, while it remains in solution when using MeOH [89]. Additional recommendations regarding this issue can be found in Päpke et al. [90] and de Boer and Wells [81]. Reported data for TBBP-A should be supported by appropriate QA/QC protocols, which can represent up to 30% of the total analytical effort. At this moment, there are no certified reference materials available for TBBP-A, although several materials have already been certified for PBDEs [91]. Such materials are necessary to evaluate the accuracy of analytical methods used for the determination of TBBP-A. Presently, the method trueness and precision can only be tested through standard addition procedures [90]. When possible, proper incubation and aging of the spiked samples should be conducted so that the spiked compounds mimic as closely as possible the behaviour of the naturally occurring analytes. Since 1999, several international interlaboratory studies have been organized with the aim of improving the quality of the analysis of BFRs [81]. A wide range of matrices, including fish and marine mammal tissue, fish oil, shellfish, sediments, sewage sludge, human milk and standard solutions, have been used during these exercises and this has enabled researchers to validate their methods and to implement reliable QC procedures. Besides PBDEs and HBCDs, the analysis of TBBP-A was also suggested, but unfortunately, only very few laboratories have submitted results for TBBP-A and/or diMe-TBBP-A. Therefore, the TBBP-A data presented in this review should presumably associated with an uncertain analytical error. 2.7. Analytical methods for TBBP-A derivatives Regarding TBBP-A derivatives, almost no data have been published to date. In a study on degradation products of TBBP-A formed after UV-exposure, ESI-MS did not prove to be very useful, while the ionization of TBBP-A degradation products by means of APPI was very efficient [83]. A potential drawback of APPI is its susceptibility towards the mobile phase composition. Higher signals were seen towards the end of the gradient elution, close to 100% acetonitrile. This particular feature could constitute a limitation for the quantitative analysis of mixtures of BFRs, their degradation products or metabolites [81].

357

Köppen et al. [85] analyzed TBBP-A bis(2,3-dibromopropylether) (TBBP-A-dbpe) in sediment and sewage sludge. Different extraction methods were compared; both fluidised bed extraction and ASE were found suitable with the latter preferred based upon its speed and lower solvent consumption. Chromatographic resolution of the extract was achieved using a Zorbax XDB-C18 column (150 mm × 4.6 mm i.d., 5 ␮m). Detection by ITD-MS using an APCI interface was evaluated, but this technique did not yield sufficient ionization, making it inadequate for quantitative analysis. Detection by DAD (220 nm) led to LODs of 10 and 22 ng/g in sediment and sewage sludge, respectively [85]. 3. Environmental levels (TBBP-A and derivatives) 3.1. Abiotic matrices Despite its extensive use, TBBP-A data for abiotic matrices are not so abundant than data available for other BFRs, such as PBDEs (Table 4). 3.1.1. Air Elevated indoor air concentrations (several orders of magnitude above those found in outdoor air) have been reported for specific occupational environments, such as electronics dismantling plants. In Sweden, TBBP-A was detected (mean 29.7 ng/m3 ) in 12 air samples at a plant for the recycling of electronics (dismantling of computers, TV sets, etc.) and in offices equipped with computers [92]. In four offices equipped with computers, the mean concentration was 0.035 ng/m3 . TBBP-A could not be detected in two outdoor samples. In another Swedish study, Sjödin et al. [93] reported a mean TBBP-A concentration of 0.036 ng/m3 in six office microenvironments containing computers, 0.093 ng/m3 from two teaching halls and 0.035 ng/m3 from two computer repair facilities. Concentrations were below detection limits (unreported) in outdoor air, indicating indoor sources of TBBP-A. They also reported high TBBPA concentrations in air from an electronics recycling plant (mean 30 ng/m3 in the dismantling hall and 140 ng/m3 in the shredder). The study found that TBBP-A was present primarily in the particulate phase rather than in the vapor phase [93]. This may suggest that passive air sampling devices that sample primarily the vapor phase (e.g. PUF disk samplers) may not be appropriate for monitoring TBBP-A. The importance of electronic goods as an emission source is underlined by Tollbäck et al. [55], who reported that the TBBP-A concentration in air from a dismantling hall within a Swedish electronics recycling plant was 13.8 ng/m3 . Inoue et al. [56] reported a mean concentration of 0.2 ng/m3 in indoor air from 26 microenvironments in Japan, where TBBP-A was found above the limit of detection (0.1 ng/m3 ) in samples from 14 out of the 26 locations. Concentrations in the matched outdoor and indoor samples from two houses in Hokkaido, Japan, were 7.1 and 9 pg/m3 , respectively for the first house, and 9.5 and 16 pg/m3 for the second [94]. Xie et al. [95] investigated the presence of TBBP-A in outdoor air from a rural site in northern Germany, over the Wadden Sea and offshore in the Northeast Atlantic Ocean (Table 4). Comparable concentrations of TBBP-A were found both at the northern German site (ranging from <0.04 to 0.85 pg/m3 ) and over the Wadden Sea (ranging from 0.31 to 0.69 pg/m3 ). Concentrations of TBBP-A in the Northeast Atlantic Ocean ranged from <0.04 to 0.17 pg/m3 [93]. The latter higher concentration was present in a sample collected off the West Norwegian coast, indicating an input source from land to ocean. Interestingly, Alaee et al. [96] reported TBBP-A to be present at 70 pg/m3 in the airborne particulates collected in 2000 in the Arctic (Dunai, Russia).

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Table 4 Mean and/or range of TBBP-A concentrations in abiotic matrices Matrix

Location

TBBP-A concentration

Ref.

Air Recycling plant Computer office Outdoor air Computer office Teaching hall Computer repair facility Outdour air Dismantling hall Shredder Dismantling hall electronics recycling plant Indoor air microenvironments Outdoor air Indoor air house Outdoor air rural site Outdoor air Outdoor air Outdoor air

Sweden Sweden Sweden Sweden Sweden Sweden Sweden Sweden Sweden Sweden Japan Japan Japan Northern Germany Wadden Sea Northeast Atlantic Arctic, Russia

29.7 ng/m3 0.035 ng/m3
[92] [92] [92] [93] [93] [93] [93] [93] [93] [55] [56] [94] [94] [95] [95] [95] [96]

Indoor dust Domestic dust Pooled domestic dust Office dust Newly constructed building Dust inside computers

Hokkaido, Japan UK EU offices Michigan, USA China

490–520 ng/g 190–340 ng/g 5–47 ng/g 0.4–2 ng/g 8.9–39.6 ␮g/g

[94] [99] [100] [101] [74]

Water Land leachate from industrial waste sites Before treatment plant After treatment plant Raw leachate from landfill Treated leachate from landfill

Japan Japan Japan Japan Japan

0.3–540 ng/L 130 ng/L 7 ng/L
[52] [52] [52] [102] [102]

Soil Outside production plant Soil Near garbage discharge site

China China China

0.12 ng/g 25.2 ± 2.7 ng/g 1.4–1.8 ␮g/g

[61] [103] [74]

Sediment Sediment Sediment downstream plastic factory

Neya River, Japan Sweden

[43] [44]

[57] [57] [57] [57] [57] [57] [57] [105] [106] [107] [58] [108] [109] [109]

Sediment upstream plastic factory

Sweden

Sediment close to BFR manufacturing site Sediment Sediment Sediment Sediment Sediment Sediment Sediment Sediments

River Skerne, UK River Tees, UK Scheldt basin Western Scheldt Dutch rivers UK rivers Detroit River Lakes Mjøsa, Losna Norway Asia

20 ng/g dw 270 ng/g (TBBP-A) 1500 ng/g (diMe-TBBP-A) 34 ng/g (TBBP-A) 24 ng/g (diMe-TBBP-A) 9.8 ␮g/g dw 25 ng/g dw 0.1–67 ng/g dw 0.1–3.2 ng/g dw 0.1–6.9 ng/g dw 2–5 ng/g dw 0.6–1.84 ng/g dw 0.04–0.13 ng/g dw <0.2–1.6 ng/g

The Netherlands The Netherlands The Netherlands UK UK UK Cork, Ireland Sweden Sweden Switzerland Ontario, Canada Montreal, Canada Canada Southern Ontario, Canada

<6.9 ng/g dw 42 ng/g dw 79 ng/g dw 7.5 ng/g dw <3.9 ng/g dw 57 ng/g dw 192 ng/g dw <0.3–220 ng/g dw 32 ng/g dw 510 ng/g dw 2.1–28.3 ng/g dw 300 ng/g dw <1–46.2 ng/g dw 14.3–43.8 ng/g dw

Sewage sludge Sewage sludge (influent) Sewage sludge (effluent) Sewage sludge Sewage sludge (influent) Sewage sludge (effluent) Sewage sludge Treated sludge Sewage sludge Sewage sludge Compost and digestate Sewage sludge from wastewater treatment and pollution control plant Sewage sludge from wastewater Sewage sludge from sewage treatment plants Sewage sludge

3.1.2. Indoor dust The presence of additive BFRs, such as PBDEs and HBCDs, in indoor dust and the consequences for human exposure has been the subject of increasing recent interest. Information on the presence of TBBP-A in indoor dust is far less extensive and

[44] [57] [57] [57] [57] [57] [57] [60] [104] [52]

appears to constitute a research gap. The available data suggest that concentrations of TBBP-A are at the low end of those found for PBDEs and HBCDs [97,98]. This is consistent with the fact that TBBP-A is used primarily as a reactive FR and as such its release from treated goods is likely to be less facile than for com-

A. Covaci et al. / J. Chromatogr. A 1216 (2009) 346–363

pounds whose use pattern is largely or exclusively as additive FRs. Takigami et al. [94] reported TBBP-A concentrations of 490 and 520 ng/g in two samples of domestic dust from Hokkaido, Japan. Similar concentrations were detected by Santillo et al. [99], who reported TBBP-A to be present above the detection limit in 4 out of 10 pooled samples of UK domestic dust. Concentrations in these four samples ranged from 190 to 340 ng/g, exceeding substantially concentrations reported in an earlier study of dust in offices from the European Parliament building, where concentrations in 9 out of 16 samples where TBBP-A was detectable were between 5 and 47 ng/g [100]. More recently, Chernyak et al. [101] monitored the change in TBBP-A concentrations in indoor dust from a newly constructed building in Michigan, USA, over the period following the building’s construction, furnishing, and occupation. A continuous increase in TBBP-A concentration in dust from 0.4 to 2.0 ng/g was observed over 1 year, suggesting that the dust samples had not yet reached saturation with TBBP-A over this period. The increase in the TBBPA concentration in the dust was less dramatic than for other BFRs, such as BDE 209. Yu and Hu [74] reported concentrations of TBBP-A ranging between 18.9 and 39.6 ␮g/g in dust samples from computers in Chinese offices (n = 4). However, these results should be interpreted with care since the analysis was performed using HPLC-DAD, yielding less specific detection than with MS detection. 3.1.3. Water Suzuki and Hasegawa [52] reported TBBP-A concentrations in water ranging from 0.3 to 540 ng/L in landfill leachates from five industrial waste sites in Japan, while concentrations of TBBP-A in influent and effluent wastewater were 130 and 7.7 ng/L, respectively. In another study, TBBP-A was measured in leachate samples taken from seven Japanese landfills [102]. Concentrations of TBBPA were up to 620 ng/L for the raw leachate and up to 11 ng/L for the treated leachate. Three sites that not only had crushed material from bulk wastes, such as waste electric and electronic equipments, but also were under operation or within a year after closure, indicated a higher concentration of BFRs than the other sites. 3.1.4. Soil To date there appears to be very few reports on concentrations of TBBP-A in soil. Jin et al. [61] reported TBBP-A at 0.12 ng/g in a soil sample taken outside a TBBP-A production plant in China. Given its reported propensity for partitioning to the atmospheric particulate phase [93] and its octanol–water partition coefficient (log Kow = 5.90), one would anticipate that soil would constitute a major sink. However, this will be influenced by the rate of degradation in soil coupled to subsequent atmospheric transport and deposition. The concentration of TBBP-A in Chinese soil measured by SPE followed by HPLC–ITD-MS was 25.2 ± 2.7 ng/g (n = 4) [103]. Another Chinese study reports TBBP-A concentrations ranging between 1.4 and 1.8 ␮g/g in soil collected near a garbage discharge site [74]. 3.1.5. Sewage sludge and sediment Similar to soil, the physico-chemical properties of TBBP-A suggest that sewage sludge and sediment are important sinks. The available data support this and reflect also the release to such matrices from industrial plants that either manufacture or use TBBP-A. In 1983, Watanabe et al. [43] reported (for the first time) TBBP-A at 20 ng/g dw in sediment from the Neya River in Japan. The concentrations of TBBP-A and its dimethylated derivative in sediment were higher downstream (270 and 1500 ng/g dw) than upstream

359

(34 and 24 ng/g dw) of a plastic factory using TBBP-A in Sweden [44]. TBBP-A was also found in sewage sludge from the wastewater treatment plant of the factory [44]. Morris et al. [57] determined TBBP-A in river and estuarine sediment samples from Belgium, the Netherlands and the UK. The highest concentration of TBBP-A (9.8 ␮g/g dw) was found in freshwater sediments from the River Skerne (UK) close to a BFR manufacturing site. The mean concentration in the River Tees, further downstream, was 25 ng/g dw. The same study also provided an indication of the expected range of concentrations in locations not influenced directly by industrial emissions, reporting concentrations of TBBP-A in sediments from the Scheldt basin (0.1–67 ng/g dw), the Western Scheldt (0.1–3.2 ng/g dw), Dutch rivers (0.1–6.9 ng/g dw) and UK rivers (2–5 ng/g dw). Lower concentrations of TBBP-A were found in sediments from Norwegian lakes (0.04–0.13 ng/g dw) [104]. TBBP-A was also quantified in influents, effluents and sewage sludge from the Netherlands (mean values of 42, <6.9, and 79 ng/g dw, respectively) and the UK (mean values of 7.5, <3.9 and 57 ng/g dw, respectively) [57]. A maximum concentration of 192 ng/g dw was quantified in a secondary treated and dewatered sludge sample from Cork, Ireland [57]. These concentrations are consistent with those reported in a survey of 57 Swedish sewage sludge samples, where concentrations ranged from <0.3 to 220 ng/g dw [105]. In 50 Swedish sewage treatment plants (STPs), TBBP-A concentrations were below LOQ (not reported) in 12 samples, while the mean concentration was 32 ng/g dw [106]. Higher concentrations were found in a few sludge samples from STPs with known or suspected point sources (textile industries, extruded polystyrene production), but in some cases, the source was unknown. However, TBBP-A concentrations were significantly lower in digester sludge samples than in the raw sludge samples, indicating anaerobic biodegradation of TBBP-A. The first comprehensive study of BFRs in source-separated compost and digestate from Switzerland showed mean TBBP-A levels of 510 ng/g dw [107]. The concentrations observed were at or above the levels found in background soils, which are the main recipient of compost and digestate. Where actually applied, compost can contribute considerably to the total input of organic pollutants to the soil. In North America, Chu et al. [58] reported concentrations of TBBP-A ranging from 2.1 to 28.3 ng/g dw in sludge samples collected from wastewater treatment and pollution control plants in Ontario, Canada. In the same study, TBBP-A was detected above the detection limit of 0.05 ng/g dw in only 3 of 55 surface sediment samples from Lake Erie and in only one sample could TBBP-A be determined quantitatively, at a concentration of 0.51 ng/g dw [58]. Importantly, the authors found that TBBP-A can undergo debromination, in agreement with the debromination of TBBP-A reported in estuarine sediments [108]. TBBP-A was also detected at a concentration of 300 ng/g dw in sewage sludge produced from the wastewaters of the Montreal area [59]. TBBP-A has been reported at moderate concentrations in Canadian sludges. Lee and Peart [109] have reported a median concentration of 12.4 ng/g (range <1–46.2 ng/g dw) in sewage sludge from 34 Canadian sewage treatment plants. Quade et al. [60] have reported low concentrations of TBBP-A in sediment from the Detroit river (range 0.60–1.84 ng/g dw). Sewage sludge collected from Southern Ontario had the same range of concentrations (14.3–43.8 ng/g dw) as reported by Lee and Peart [109]. In Asia, low concentrations of TBBP-A (<0.2–1.6 ng/g) were determined in sediments sampled before treatment in water treatment plants in Japan [52].

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A. Covaci et al. / J. Chromatogr. A 1216 (2009) 346–363

Table 5 Mean and range of TBBP-A concentrations (ng/g lipid weight) in biological matrices Species

Tissue

Location

TBBP-A concentration (ng/g lw)

Refs.

Invertebrates Common whelk Sea star Sea star Hermit crab Mysid

Whole Whole Whole Whole Whole

North Sea Scheldt estuary Tees estuary North Sea Scheldt estuary

5.0–96 <1–2 205 <1–35 0.8–0.9

[57] [57] [57] [57] [110]

Fish Whiting Cod Hake Eel Eel Yellow eel Yellow eel Perch, pike, smelt, vendace, trout Atlantic cod Bull shark Atlantic sharpnose shark

Muscle Liver Liver Muscle Muscle Muscle Muscle Muscle Muscle Muscle Muscle

North Sea North Sea Atlantic Scheldt estuary Dutch rivers Scheldt basin Dutch rivers Norway Norway Florida, USA Florida, USA

<97 to 245 (mean 136) <0.3–1.8 <0.2 <0.1–13 (mean 1.6) <0.1–1.3 (mean 0.3) <0.1–2.1 <0.1–1.0 1.0–13.7 0.5 and 2.5 9.5 ± 12.0 0.87 ± 0.50

[57] [57] [57] [57] [57] [57] [57] [104] [111] [68] [68]

Birds Cormorant Common tern Peregrine falcon, White-tailed sea eagle, Osprey, Golden eagle, Peregrine falcon

Liver Egg Egg Egg

Wales and England Western Scheldt Norway Norway

2.5–14 <2.9 <0.003–0.013a nd (TBBP-A) <0.1–940 (diMe-TBBP-A)

[57] [57] [112] [113]

Marine mammals Harbour seal Harbour porpoise Harbour porpoise Harbour porpoise Harbour porpoise Bottlenose dolphin

Blubber Blubber Blubber Blubber Blubber Blubber

Wadden Sea North Sea North Sea Tyne/Tees UK Florida, USA

<14 <11 0.1–418 0.31 6–35a 1.2 ± 3.0

[57] [57] [57] [57] [62] [68]

Humans Electronics dismantling Computer technicians Electronics dismantling Circuit board producers Laboratory personnel General population General population General population

Serum Serum Serum Serum Serum Serum Serum Adipose tissue

Sweden Sweden Norway Norway Norway Norway Japan New York, USA

1.1–4.0 0.55–1.84 0.64–1.8 (mean 1.3) <0.1–0.80 (mean 0.54) <0.1–0.52 (mean 0.34) 0.34–0.71 1.35 0.048 ± 0.102

[114] [115] [116] [116] [116] [79] [117] [68]

a

Concentrations in ng/g wet weight.

3.2. Biological matrices Despite the extensive use of TBBP-A, data for biotic matrices are scarce (Table 5). Herzke et al. [112] determined TBBP-A in two eggs from each of four different Norwegian bird of prey species (osprey, golden eagle, white-tailed sea eagle and peregrine falcon), which have different feeding habits and habitats. These eggs were sampled between 1992 and 2002. TBBP-A was detected in all eight samples in a concentration range of <3–13 pg/g wet weight (ww), which indicates that TBBP-A is distributed widely in a broad range of prey items of predatory birds in Norway [112]. Law et al. [62] determined TBBP-A in the blubber of 68 porpoises (Phocoena phocoena) stranded in UK waters between 1994 and 2003. TBBP-A was detected in only 18 samples, with concentrations between 6 and 35 ng/g ww. Morris et al. [57] determined TBBP-A in a variety of aquatic biota from the North Sea, including five cormorant (Phalacrocorax carbo) liver samples from England. Levels ranged from 2.5–14 ng/g lipid weight (lw) [57]. TBBP-A was also detected in mysid shrimp (Neomysis integer) from two sites in the Scheldt estuary at concentrations of 0.8 and 0.9 ng/g lw [110]. In liver of Atlantic cod (Gadus morhua) from the Norwegian Arctic collected in 1998 and 2002, TBBP-A concentrations ranged from 0.5 to 2.5 ng/g lw [111].

Recently, Johnson-Restrepo et al. [68] have measured the concentrations of TBBP-A in three marine top-predators from coastal waters of Florida, USA. The overall mean concentrations (mean ± SD) of TBBP-A were 1.2 ± 3.0 ng/g lw, 9.5 ± 12.0 ng/g lw, and 0.87 ± 0.50 ng/g lw in bottlenose dolphin blubber (n = 15), bull shark muscle (n = 13) and Atlantic sharpnose shark muscle (n = 3), respectively. The highest concentration of TBBP-A (35.6 ng/g lw) was measured in bull shark muscle. Thirty-three peregrine falcon eggs from South Greenland were analyzed by Vorkamp et al. [113]. TBBP-A could not be detected in any of the eggs whereas diMe-TBBP-A was quantifiable in 29 out of 33 eggs. Concentrations of diMe-TBBP-A ranged between <0.1 and 940 ng/g lw, with a mean concentration of 280 ng/g lw. In a preliminary recent screening study, TBBP-A and diMe-TBBP-A were not detected in any of the samples of egg, liver and adipose tissue of marine biota from Greenland and the Faroe Islands, indicating limited or no transport of these compounds to remote areas [86,118]. 3.3. Food To the best of our knowledge, dedicated studies that focus on TBBP-A and its derivatives in foodstuffs have not been reported to date. It is assumed that less-persistent BFRs, such as TBBP-A, do

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not biomagnify and hence will not be present at significant levels in food of animal origin [119]. Saito et al. [120] developed a method for the determination of TBBP-A in cattle adipose tissue and swine organs, but this was not applied to real samples. Only a limited number of studies have investigated the presence of TBBP-A in animal species (e.g. fish) that are suitable for human consumption [57,104]. These studies both reported that concentrations of TBBP-A were low (Section 3.2 and Table 4). Recently, 19 composite food group samples have been analyzed in the UK for a wide range of pollutants, including TBBP-A [121]. The study did not detect any TBBP-A at concentrations above the LOD (0.36 ng/g). The UK Food Standards Agency estimated the dietary intake from fisheries products (composite samples of 48 species of farmed and wild fish and shellfish, together with ten samples of fish oil dietary supplements) through the analysis of BFRs, including TBBP-A [122]. TBBP-A was not detected in any of the samples tested. Based on these results, the UK Committee on Toxicity of Chemicals in Food, Consumer Products and the Environment concluded that the levels of TBBP-A detected in fish and shellfish do not raise toxicological concerns and that the estimated dietary exposure to TBBP-A (<1.6 ng/kg body weight/day) seems to have limited implications for health [123]. No data are available for TBBP-A derivatives. Despite the low levels of TBBP-A reported in food for human consumption, the European Food Safety Authority [124] issued advice in 2006 concerning the routine measurements of BFRs in food and feed. A scientific committee concluded that, based on literature data, TBBP-A should be included within the monitoring programs [124]. This conclusion was reached after considering the following criteria: (i) analytical feasibility to measure the chemical compounds routinely in accredited laboratories, (ii) production volume, (iii) occurrence of the chemical compounds in food and feed, (iv) persistence and (v) toxicity. In contrast to what was decided for the PBDEs, HBCDs and PBBs, TBBP-A was not recommended directly for inclusion in the core group of analytes to be monitored. The scientific panel pointed out that it would be desirable to initiate a specific research program for reactive BFRs, such as TBBP-A.

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Thomsen et al. [116] analyzed serum from Norwegian individuals working at an electronics dismantling facility, in the production of printed circuit boards and as laboratory personnel, the latter being a control group. In that study, TBBP-A levels were elevated significantly (p < 0.05) in the dismantlers (1.3 ng/g lw), while in both other groups the levels were lower (0.54 and 0.34 ng/g lw, respectively). Data on TBBP-A in non-occupationally exposed individuals was published by Thomsen et al. [79], where time trends from 1977 to 1999 were also investigated. TBBP-A could not be found in the oldest serum pools (1977 and 1981), but was present in all other samples. Further, Thomsen et al. [79] also looked at age classes of the 1998population. TBBP-A levels tended to be the highest in the age group of 0–4 years. This was also the only age group where diMe-TBBP-A was found, although at a level close to the LOQ. Further, no concentration vs. age relationship could be observed. Concentrations in that study ranged between 0.34–0.71 ng/g lw. After further method optimization, the same samples were re-analyzed and Me-TBBP-A was found in all samples [80]. In Japan, TBBP-A was analyzed in 24 blood samples from adult volunteers. TBBP-A was detected in only 14 of these samples with a mean concentration of 1.35 ng/g lw [117]. Recently, Johnson-Restrepo et al. [68] measured the concentrations of TBBP-A in 20 adipose tissue samples from New York, USA. The overall mean concentration (mean ± SD) of TBBP-A was 0.048 ± 0.102 ng/g lw, with a maximum concentration of 0.46 ng/g lw. TBBP-A correlated well with concentrations of HBCDs, but not with those of PBDEs. Moreover, concentrations of TBBP-A were 10fold lower than HBCD concentrations and 3–4 orders of magnitude lower than PBDEs measured in the same samples. Detection of TBBP-A in humans can be hampered by the short biological half-life of the compound, which has been estimated to be 2 days [89,114]. This is not surprising since TBBP-A is a phenol that can be rapidly conjugated and subsequently excreted [125]. Still, TBBP-A may accumulate in humans, but a continuous exposure to this BFR is required to maintain a detectable level in the human subject.

3.4. Humans 4. Concluding remarks In general, reports of TBBP-A in human samples are scarce. The first report of TBBP-A being present in human tissues dates back to 1979. In Arkansas, USA, TBBP-A was found in human hair in the vicinity of TBBP-A manufacturing sites [44]. More recent reports of TBBP-A in human samples are also mainly focused on occupationally exposed workers [114–116], because of the greater likelihood of exposure in occupational environments (Table 5). Due to its limited presence in foodstuffs (as far as it has been investigated), direct exposure via inhalation might be considered the predominant route of human exposure [119]. In this respect, occupationally exposed workers are at higher risk than the general population. High concentrations of TBBP-A in the air (30 ng/m3 ) inside an electronics dismantling area of a dismantling facility were measured [119]. In a follow-up study of the people working in the same electronics recycling facility, Hagmar et al. [114] revealed the presence of TBBP-A (range 1.1–4.0 ng/g lw) in the serum of the workers engaged in the recycling process, which indicates systemic uptake of this chemical. Jakobsson et al. [115] investigated TBBP-A exposure in computer technicians. TBBP-A could be found in 80% of the technicians, while the compound could not be measured in the serum of a control group comprising office clerks and hospital cleaners. The concentrations found in this study ranged between <0.55 and 1.84 ng/g lw, which are comparable to the concentrations reported by Hagmar et al. [114].

The LC–MS/MS method appears to be the method of choice for TBBP-A analysis because no derivatization is required. Moreover, the use of 13 C-labelled TBBP-A as an internal standard enhances the quality of the analytical data through compensation for matrixrelated effects that can affect analyte ion intensity, trueness and reproducibility. Since TBBP-A is a reactive BFR, its release from treated goods is much less pronounced than for additive BFRs, such as HBCDs and PBDEs. Consequently, this is reflected in the low concentrations reported in indoor dust, air and food stuffs, which makes human exposure to TBBP-A via air inhalation, dust ingestion and diet far less significant than the estimated intake for additive BFRs, such as HBCDs. This review also points that there are still knowledge gaps regarding the presence of TBBP-A and its derivatives in indoor and outdoor air, in indoor dust and food, as well as regarding human exposure via these pathways. The possibility of degradation in soil and sediment, debromination and the possible pathways thereof, together with the factors affecting these processes, also requires further investigation. Moreover, future work should also focus on environmental fate and the human exposure around production or usage sites, primarily in Asia. Such sites are seen as the worstcase scenarios for occupational or accidental exposure, while the exposure of the general population seems to be very low given the low leaching capacity of TBBP-A from finished products.

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Acknowledgements The authors would like to thank Steve Dungey of the UK Environment Agency for supplying a copy of the draft EU risk assessment (environment) document for TBBP-A. Adrian Covaci acknowledges a postdoctoral fellowship of the Funds of Scientific Research Flanders (FWO), while Tinne Geens was funded by the Research Funds of the University of Antwerp. UK work (Robin Law) has been funded by the Department for Environment, Food and Rural Affairs (Defra). Mohamed Abdallah acknowledges gratefully the provision of studentships from the Egyptian Government and the Egyptian Ministry of Higher Education. References [1] European Risk Assessment Report on 2,2 ,6,6 -tetrabromo-4,4 isopropylidenediphenol(tetrabromobisphenol-A or TBBP-A). Part I, Environment, European Commission, Joint Research Centre, European Chemicals Bureau, in draft, 2008. [2] European Risk Assessment Report on 2,2 ,6,6 -tetrabromo-4,4 isopropylidenediphenol(tetrabromobisphenol-A or TBBP-A). Part II, Human health, vol. 63, European Commission, Joint Research Centre, European Chemicals Bureau, EUR22161E, 2006. [3] Environmental Health Criteria 172, Tetrabromobisphenol A and Derivatives, International Programme on Chemical Safety, World Health Organization, Geneva, 1995. [4] R.J. Law, C.R. Allchin, J. de Boer, A. Covaci, D. Herzke, P. Lepom, S. Morris, J. Tronczynski, C.A. de Wit, Chemosphere 64 (2006) 187. [5] Bromine Science Environmental Forum, http://www.bsef.com (accessed December 15, 2007). [6] H. Hakk, A Survey of Tetrabromobisphenol A. Second International Workshop on Brominated Flame Retardants, BFR 2001, Stockholm University, Sweden, 2001. [7] P.A. Arias, Presented at the Second International Workshop on Brominated Flame Retardants BFR 2001, Stockholm, Sweden, May 14–16, 2001, p. 17. [8] EBFRIP questions Norwegian proposal to restrict the use of TBBP-A and HBCD in consumer products, European Brominated Flame Retardant Industry Panel (EBFRIP), statement June 4, 2007. [9] Flame Retardant Fact Sheet—Tetrabromobisphenol A (TBBP-A), European Flame Retardants Association (EFRA), http://www.cefic-efra.com (accessed December 15, 2007). [10] Brominated flame retardants, Environmental Project 494, Danish Environmental Protection Agency, 1999. [11] Directive 2002/95/EC of the European Parliament and of the Council of January 27, 2003 on the restriction of the use of certain hazardous substances in electrical and electronic equipment, Off. J. Eur. Comm. L 37 (2003) 19. [12] G.W. Gribble, Environ. Sci. Pollut. Res. 7 (2000) 37. [13] Directive 2002/96/EC of the European Parliament and of the Council of January 27, 2003 on waste electrical and electronic equipment (WEEE), Off. J. Eur. Comm. L 037 (2003) 24. [14] Regulation 2364/2000/EC of October 25, 2000 concerning the fourth list of priority substances as foreseen under Council Regulation (ECC) No. 793/93, Off. J. Eur. Comm. L 273 (2000) 5. [15] Regulation (EC) No. 1907/2006 of the European Parliament and of the Council of December 18, 2006 concerning the Registration, Evaluation, Authorisation and Restriction of Chemicals (REACH), establishing a European Chemicals Agency, amending Directive 1999/45/EC and repealing Council Regulation (EEC) No. 793/93 and Commission Regulation (EC) No. 1488/94 as well as Council Directive 76/769/EEC and Commission Directives 91/155/EEC, 93/67/EEC, 93/105/EC and 2000/21/EC, Off. J. Eur. Comm. L 396 (2006) 1. [16] S. Eisenreich, S. Munn, S. Pakalin, Proceedings of the 4th International Workshop on Brominated Flame Retardants, Amsterdam, The Netherlands, April 24-27, 2007 (unpaginated). [17] Draft updated risk assessment of 4,4 -isopropylidenediphenol (bisphenol-A), European Chemicals Bureau (ECB), 2007, R325 0705 env1. [18] Decision 2455/2001/EC of November 20, 2001. 1, Establishing the list of priority substances in the field of water policy and amending Directive 2000/60/EC, Off. J. Eur. Comm. L 331 (2001) 1. [19] K. Akutsu, M. Kitagawa, H. Nakazawa, T. Makino, K. Iwazaki, H. Oda, S. Hori, Chemosphere 53 (2003) 645. [20] Y. Yoshida, T. Kawamura, K. Yamamoto, H. Sekii, S. Komoto, Organohalogen Compd. 67 (2005) 2127. [21] M.H. Wong, S.C. Wu, W.J. Deng, X.Z. Yu, Q. Luo, A.O.W. Leung, C.S.C. Wong, W.J. Luksemburg, A.S. Wong, Environ. Pollut. 149 (2007) 131. [22] D. Gutzman, R. Chenier, J. Pasternak, L. Suffredine, K. Taylor, Proceedings of the Third International Workshop on Brominated Flame Retardants, BFR 2004, Toronto, ON, Canada, 2004. [23] S. Kitamura, N. Jinno, S. Ohta, H. Kuroki, N. Fujimoto, Biochem. Biophys. Res. Commun. 293 (2002) 554. [24] M. Ghisari, E.C. Bonefeld-Jorgensen, Mol. Cell. Endocrinol. 244 (2005) 31.

[25] M.H. Kester, S. Bulduk, H. van Toor, D. Tibboel, W. Meinl, H. Glatt, C.N. Falany, M.W. Coughtrie, A.G. Schuur, A. Brouwer, T.J. Visser, J. Clin. Endocrinol. Metab. 87 (2002) 1142. [26] I.A. Meerts, R.J. Letcher, S. Hoving, G. Marsh, A.A. Bergman, J.G. Lemmen, B. van Der Burg, A. Brouwer, Environ. Health Perspect. 109 (2001) 399. [27] M. Samuelsen, C. Olsen, J.A. Holme, E. Meussen-Elholm, A. Bergmann, J.K. Hongslo, Cell Biol. Toxicol. 17 (2001) 139. [28] C.M. Olsen, E. Meussen-Elholm, M. Samuelsen, J.A. Holme, J.K. Hongslo, Pharmacol. Toxicol. 92 (2003) 180. [29] S. Kitamura, T. Kato, M. Iida, N. Jinno, T. Suzuki, S. Ohta, N. Fujimoto, H. Hanada, K. Kashiwagi, A. Kashiwagi, Life Sci. 76 (2005) 1589. [30] I.A.T.M. Meerts, J.J. van Zanden, E.A.C. Luijks, I. van Leeuwen-Bol, G. Marsh, E. Jakobsson, Å. Bergman, A. Brouwer, Toxicol. Sci. 56 (2000) 95. [31] T. Hamers, J.H. Kamstra, E. Sonneveld, A.J. Murk, M.H.A. Kester, P.L. Andersson, J. Legler, A. Brouwer, Toxicol. Sci. 92 (2006) 157. [32] E. Mariussen, F. Fonnum, Neurochem. Int. 43 (2003) 533. [33] S. Pullen, R. Boecker, G. Tiegs, Toxicology 184 (2003) 11. [34] T. Reistad, E. Mariussen, F. Fonnum, Toxicol. Sci. 83 (2005) 89. [35] S. Strack, T. Detzel, M. Wahl, B. Kuch, H.F. Krug, Chemosphere 67 (2007) S405. [36] R.V. Kuiper, E.J. van den Brandhof, P.E.G. Leonards, L.T.M. van der Ven, P.W. Wester, J.G. Vos, Arch. Toxicol. 81 (2007) 1. [37] Draft Risk Reduction Monograph No. 3, Brominated Flame Retardants, Organisation for Economic Co-operation and Development, May 1994. [38] M. Ash, I. Ash, The Index of Flame Retardants, Gower Publishing Limited, Aldershot, UK, 1997, ISBN 0-566-07885-6. [39] A.-M. Prins, C. Doumen, J. Kaspersma, Proceedings of the Flame Retardants 2000 Conference, London, February 8–9, 2000, p. 77. [40] Y. Bar Yaakov, L. Utevski, J. Reyes, P. Georlette, S. Bron, J.M. Lopez-Cuesta, Proceedings of the Flame Retardants 2000 Conference, London, February 8–9, 2000, p. 87. [41] D. De Schryver, T. DeSoto, R. Dawson, S.D. Landry, R. Herbiet, Proceedings of the Flame Retardants 2002 Conference, Interscience Communications Limited, London, 2002. [42] B. Plaitin, A. Fonzé, R. Braibont, Proceedings of the Flame Retardants ‘98 Conference, London, February 3–4, 1998, p. 139. [43] I. Watanabe, T. Kashimoto, R. Tatsukawa, Bull. Environ. Contam. Toxicol 31 (1983) 48. [44] U. Sellström, B. Jansson, Chemosphere 31 (1995) 3085. [45] J. de Boer, C. Allchin, B. Zegers, J.P. Boon, S.H. Brandsma, S. Morris, A.W. Kruijt, I. van der Steen, J.M. van Hesselingen, J.J.H. Haftka, HBCD and TBBP-A in sewage sludge, sediments and biota, including interlaboratory study, RIVO Report No. C033/02, September 2002. [46] A.S. Allard, M. Remberger, A.H. Neilson, Appl. Environ. Microbiol. 53 (1987) 839. [47] A. Covaci, S. Voorspoels, J. de Boer, Environ. Int. 29 (2003) 735. [48] A. Covaci, S. Voorspoels, L. Ramos, H. Neels, R. Blust, Chromatogr. A 1153 (2007) 145. [49] E. Eljarrat, D. Barcelo, Trends Anal. Chem. 23 (2004) 727. [50] R.A. Hites, Environ. Sci. Technol. 38 (2004) 945. [51] A. Covaci, A.C. Gerecke, R.J. Law, M. Kohler, N.V. Heeb, H. Leslie, C.R. Allchin, J. de Boer, Environ. Sci. Technol. 40 (2006) 3679. [52] S. Suzuki, A. Hasegawa, Anal. Sci. 22 (2006) 469. [53] H. Gallart-Ayala, E. Moyano, M.T. Galceran, Rapid Commun. Mass Spectrom. 21 (2007) 4039. [54] H. Sambe, K. Hoshina, K. Hosoya, J. Haginaka, J. Chromatogr. A 1134 (2006) 16. [55] J. Tollbäck, C. Crescenzi, E. Dyremark, J. Chromatogr. A 1104 (2006) 106. [56] K. Inoue, S. Yoshida, S. Nakayama, R. Ito, N. Okanouchi, H. Nakazawa, Arch. Environ. Contam. Toxicol. 51 (2006) 503. [57] S. Morris, C.R. Allchin, B.N. Zegers, J.J.H. Haftka, J.P. Boon, C. Belpaire, P.E.G. Leonards, S.P.J. van Leeuwen, J. de Boer, Environ. Sci. Technol. 38 (2004) 5497. [58] S. Chu, G.D. Haffner, R.J. Letcher, J. Chromatogr. A 1097 (2005) 25. [59] R. Saint-Louis, E. Pelletier, Analyst 129 (2004) 724. [60] S.C. Quade, M. Alaee, C. Marvin, R. Hale, K.R. Solomon, N.J. Bunce, A.T. Fisk, Organohalogen Compd. 62 (2003) 327. [61] J. Jin, H. Peng, Y. Wang, R. Yang, J. Cui, Organohalogen Compd. 68 (2006) 85. [62] R.J. Law, P. Bersuder, C. Allchin, J. Barry, Environ. Sci. Technol. 40 (2006) 2177. [63] S. Morris, P. Bersuder, C.R. Allchin, B. Zegers, J.P. Boon, P.E.G. Leonards, J. de Boer, Trends Anal. Chem. 25 (2006) 343. [64] K. Granby, T. S Cederberg, Proceedings of the Fourth International Workshop on Brominated Flame Retardants BFR 2007, Amsterdam, The Netherlands, 2007. [65] U. Berger, D. Herzke, T.M. Sandanger, Anal. Chem. 76 (2004) 441. [66] R. Cariou, J.P. Antignac, P. Marchand, A. Berrebi, D. Zalko, F. Andre, B. Le Bizec, J. Chromatogr. A 1100 (2005) 144. [67] T. Hayama, H. Yoshida, S. Onimaru, S. Yonekura, H. Kuroki, K. Todoroki, H. Nohta, M. Yamaguchi, J. Chromatogr. B 809 (2004) 131. [68] B. Johnson-Restrepo, D. Adams, K. Kannan, Chemosphere 70 (2008) 1935. [69] M. Schlummer, F. Brandl, A. Mäurer, R. van Eldik, J. Chromatogr. A 1064 (2005) 39. [70] M. Pöhlein, A. Segura Llopis, M. Wolf, R. van Eldik, J. Chromatogr. A 1066 (2005) 111. [71] J. Llorca-Porcel, G. Martinez-Sanchez, B. Alvarez, M.A. Cobollo, I. Valor, Anal. Chim. Acta 569 (2006) 113. [72] M. Polo, M. Llompart, C. Garcia-Jares, G. Gomez-Noya, M.H. Bollain, R. Cela, J. Chromatogr. A 1124 (2006) 11.

A. Covaci et al. / J. Chromatogr. A 1216 (2009) 346–363 [73] [74] [75] [76] [77] [78] [79] [80] [81] [82] [83] [84] [85] [86] [87] [88] [89]

[90] [91] [92] [93] [94] [95] [96] [97] [98] [99]

[100]

E. Blanco, M.C. Casais, M.C. Mejuto, R. Cela, Anal. Chem. 78 (2006) 2772. C. Yu, B. Hu, J. Chromatogr. A 1160 (2007) 71. A.M. Altwaiq, M. Wolf, R. van Eldik, Anal. Chim. Acta 491 (2003) 111. S. Kikuchi, K. Kawauchi, S. Ooki, M. Kurosawa, H. Honjo, T. Yagishita, Anal. Sci. 20 (2004) 1111. F.T. Dettmer, H. Wichmann, J. de Boer, M. Bahadir, Chemosphere 39 (1999) 1523. A.S. Vasquez, A. Martin, J.M. Costa-Fernandez, J.R. Encinar, N. Bordel, R. Pereiro, A. Sanz-Mendel, Anal. Bioanal. Chem. 389 (2007) 683. C. Thomsen, E. Lundanes, G. Becher, Environ. Sci. Technol. 36 (2002) 1414. C. Thomsen, V. Horpestad Liane, G. Becher, J. Chromatogr. B 846 (2007) 252. J. de Boer, D.E. Wells, Trends Anal. Chem. 25 (2006) 364. P. Korytár, A. Covaci, P.E.G. Leonards, J. de Boer, U.A.Th. Brinkman, J. Chromatogr. A 1100 (2005) 20. L. Debrauwer, A. Riu, M. Jouahri, E. Rathahao, I. Jouanin, J.P. Antignac, R. Cariou, B. Le Bizec, D. Zalko, J. Chromatogr. A 1082 (2005) 98. K. Worrall, P. Hancock, A. Fernandes, M. Driffield, Organohalogen Compd. 69 (2007) 698. R. Köppen, R. Becker, C. Jung, C. Piechotta, I. Nehls, Anal. Bioanal. Chem. 384 (2006) 1485. M. Frederiksen, K. Vorkamp, R. Bossi, F. Riget, M. Dam, B. Svensmark, Int. J. Environ. Anal. Chem. 87 (2007) 1095. D. Zalko, C. Prouillac, A. Riu, E. Perdu, L. Dolo, I. Jouanin, C. Canlet, L. Debrauwer, J.P. Cravedi, Chemosphere 64 (2006) 318. E. Blanco, M.C. Casais, M.C. Mejuto, R. Cela, J. Chromatogr. A 1071 (2005) 205. A. Sjödin, Occupational and dietary exposure to organohalogenated substances with special emphasis on polybrominated flame retardants, Ph.D. Thesis, Stockholm University, Sweden, 2000. O. Päpke, P. Fürst, T. Herrmann, Talanta 63 (2004) 1203. H.M. Stapleton, J.M. Keller, M.M. Schantz, J.R. Kucklick, S.D. Leigh, S.A. Wise, Anal. Bioanal. Chem. 387 (2007) 2365. Å. Bergman, M. Athanasiadou, E. Klasson Wehler, A. Sjödin, Organohalogen Compd. 43 (1999) 89. A. Sjödin, H. Carlsson, K. Thuresson, S. Sjolin, A. Bergman, C. Ostman, Environ. Sci. Technol. 35 (2001) 448. H. Takigami, G. Suzuki, Y. Hirai, S. Sakai, Organohalogen Compd. 69 (2007) 2785. Z. Xie, R. Ebinghaus, R. Lohmann, O. Heemken, A. Cabaa, W. Puttmann, Anal. Chim. Acta 584 (2007) 333. M. Alaee, D. Muir, C. Cannon, P. Helm, T. Harner, T. Bidleman, Contam. Assess. 1 Rep. 2 (2003) 116. M. Abdallah, S. Harrad, C. Ibarra, M. Diamond, L. Melymuk, M. Robson, A. Covaci, Environ. Sci. Technol. 42 (2008) 459. S. Harrad, C. Ibarra, M. Diamond, L. Melymuk, M. Robson, J. Douwes, L. Roosens, A.C. Dirtu, A. Covaci, Environ. Int. 34 (2008) 232. D. Santillo, I. Labunska, H. Davidson, P. Johnston, M. Strutt, O. Knowles, Greenpeace Research Laboratories technical note 01/2003 (GRL-TN-012003) 2003, http://www.greenpeace.to/publications/housedust uk 2003.pdf (accessed January 4th 2008). P.E.G. Leonards, D. Santillo, K. Brigden, I. van der Veen, J. von Hesselingen, J. de Boer, P. Johnston, Proceedings of the Second International Workshop on Brominated Flame Retardants BFR 2001, Stockholm, Sweden, May 14–16, 2001, p. 299.

363

[101] S. Chernyak, S. Batterman, C. Godwin, C. Jia, S. Charles, Organohalogen Compd. 69 (2007) 994. [102] M. Osako, Y.J. Kim, S.I. Sakai, Chemosphere 57 (2004) 1571. [103] H. Peng, J. Jin, Y. Wang, W.Z. Liu, R.M. Yang, Chin. J. Anal. Chem. 35 (2007) 549. [104] M. Schlabach, E. Fjeld, H. Gundersen, E. Mariussen, G. Kjellberg, E. Breivik, Organohalogen Compd. 66 (2004) 3730. [105] K. Öberg, K. Warman, T. Oberg, Chemosphere 48 (2002) 805. [106] C. de Wit, K. Nylund, U. Eriksson, A. Kierkegaard, L. Asplund, Organohalogen Compd. 69 (2007) 2686. [107] R.C. Brandli, T. Kupper, T.D. Bucheli, M. Zennegg, S. Huber, D. Ortelli, J. Muller, C. Schaffner, S. Iozza, P. Schmid, U. Berger, P. Edder, M. Oehme, F.X. Stadelman, J. Tarradellas, J. Environ. Monit. 9 (2007) 465. [108] J. Voordeckers, D. Fennell, K. Jones, M. Haggblom, Environ. Sci. Technol. 36 (2002) 696. [109] H.B. Lee, T.E. Peart, Water Qual. Res. J. Canada 37 (2002) 681. [110] T.A. Verslycke, A.D. Vethaak, K. Arijs, C.R. Janssen, Environ. Pollut. 136 (2005) 19. [111] E. Fjeld, M. Schlabach, J.A. Berge, T. Eggen, P. Snilsberg, G. Källberg, S. Rognerud, E.K. Enge, A. Borgen, H. Gundersen, Kartlegging av utvalgte nye organiske miljøgifter-bromerte flammehemmere, klorerte parafiner, bisfenol A og trichlosan. Norsk institutt för vannforskning (NIVA), Rapport 48092004, Oslo, Norway (in Norwegian), 2004 (available from http://www. nilu.no). [112] D. Herzke, U. Berger, R. Kallenborn, T. Nygard, W. Vetter, Chemosphere 61 (2005) 441. [113] K. Vorkamp, M. Thomsen, K. Falk, H. Leslie, S. Møller, P.B. Sørensen, Environ. Sci. Technol. 39 (2005) 8199. [114] L. Hagmar, A. Sjödin, P. Häglund, K. Thuresson, L. Rylander, Å. Bergman, Organohalogen Compd. 47 (2000) 198. [115] K. Jakobsson, K. Thuresson, L. Rylander, A. Sjödin, L. Hagmar, Å. Bergman, Chemosphere 46 (2002) 709. [116] C. Thomsen, E. Lundanes, G. Becher, J. Environ. Monit. 23 (2001) 366. [117] J. Nagayama, H. Tsuji, T. Takasuga, Organohalogen Compd. 48 (2000) 27. [118] M. Frederiksen, K. Vorkamp, R. Bossi, B. Svensmark, Proceedings of the Fourth International Workshop on Brominated Flame Retardants, BFR 2007, Amsterdam, The Netherlands, 2007. [119] A. Sjödin D.G.Jr., Å. Patterson, Bergman, Environ. Int. 29 (2003) 829. [120] K. Saito, A. Sjödin, C.D. Sandau, M.D. Davis, H. Nakazawa, Y. Matsuki, D.G. Patterson Jr., Chemosphere 57 (2004) 373. [121] Food Standards Agency, Brominated chemicals: UK dietary intakes, Food Survey Information Sheet 10/06, 2006a, 26 pp. (http://www.food.gov.uk/ multimedia/pdfs/fsis1006.pdf). [122] Food Standards Agency, Brominated chemicals in farmed, wild fish, shellfish and fish oil dietary supplements, Food Survey Information Sheet 04/06, 2006b, 32 pp., http://www.food.gov.uk/multimedia/pdfs/fsis0406.pdf. [123] UK Committee on Toxicity of Chemicals in Food, Consumer products and the Environment, http://www.food.gov.uk/multimedia/pdfs/fsis1006.pdf. [124] Advice of the scientific panel on contamination in the food chain on a request from the Commission related to relevant chemical compounds in the group of brominated flame retardants for monitoring in feed and food, EFSA J. 328 (2006) 1. [125] H. Hakk, G. Larsen, Å. Bergman, U. Örn, Xenobiotica 30 (2000) 881.