RAPPORT 4697

REPORT 5065

REPORT 5065

Brom

inated Flame R

etardantsR

EP

OR

T 5

06

5

POP

POP

POPpersistent organic pollutants

isbn 91-620-5065-6issn 0282-7298

swedish environmental protection agency

Brominated

Brominated

Brominated Flame RetardantsFlame Retardants

Flame Retardants

Flame Retardants

The fact that several brominated flame retardants (BFRs)have been found in the environment, sometimes at increasinglevels, has caused great concern in many fora.

The time trends studied, indicate increased levels of manyBFRs in the environment since the 1970s but the levels ofimportant lower-brominated compounds have begun todecline in the Baltic Sea since voluntary withdrawal of use ina number of countries. However, the human breast milktrend indicates that levels are increasing exponentially,doubling every five years.

The results may indicate that humans are exposed to thesesubstances not just from the diet, but also from exposure toelectronic appliances and textiles.

These findings, and many more, are discussed in this report.The results indicate that BFRs may be a new "PCB problem".

Brominated

Brominated

Brominated

Flame Retardants

Flame Retardants

Flame Retardants

Cynthia A. de Wit

RAPPORT 5065 00-07-03, 10.331

BrominatedFlame Retardants

Cynthia A. de Wit

swedish environmental protection agency

Contact: Niklas Johansson

Telephone: +46 8 698 14 38

E-mail: Niklas.Johansson@environ.se

The author assumes sole responsibility for the contentsof this report, which, therefore, cannot be cited as representing,

the viewpoint of the Swedish Environmental Protection Agency.The report has been submitted to external referees for review.

The author’s address:Institute of Applied Environmental Research (ITM)

Stockholm UniversitySE-106 91 Stockholm, Sweden

E-mail: cynthia.de.wit@itm.su.se

Address for orders:Swedish Environmental Protection Agency

Customer ServicesSE-106 48 Stockholm, Sweden

Telephone: +46 8 698 12 00Fax: +46 8 698 15 15

E-mail: kundtjanst@environ.seInternet: www.environ.se

Bookstore: www.miljobokhandeln.com

ISBN 91-620-5065-6ISSN 0282-7298

© Swedish Environmental Protection AgencyPrinted by: Elanders Gotab, Stockholm, Sweden, 2000

Edition: 800 copies

Preface

The results presented in this report are, in many cases,from Swedish research projects funded by a number ofnational and international research programmes. Theseinclude grants from the Swedish Natural Science ResearchCouncil, the Swedish Medical Research Council, the Eu-ropean Commission Environment and Climate Program,the Swedish Environmental Protection Agency’s monitor-ing program (Brominated Flame Retardant Project) andthe Persistent Organic Pollutants (POP) research programfunded by the Swedish Environmental Protection Agencyfrom 1992 to 1998.

This review has been compiled by Cynthia A. de Wit,Stockholm University, at the request of the steering groupof the POP research program.

Swedish Environmental Protection Agency

Stockholm, Sweden 2000

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Contents

Abbreviations. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7Summary in Swedish . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 81. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 111.1 What are flame retardants?. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 111.2 Which substances are covered in this report? . . . . . . . . . . . . . . . . . . . . . . 121.2.1 Polybrominated diphenyl ethers (PBDEs) . . . . . . . . . . . . . . . . . . . . . . . 131.2.2 Tetrabromobisphenol A (TBBPA) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 151.2.3 Hexabromocyclododecane (HBCD) . . . . . . . . . . . . . . . . . . . . . . . . . . . . 151.2.4 Polybrominated biphenyls (PBBs) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16

2. Brominated flame retardant chemistry . . . . . . . . . . . . . . . 172.1 Synthesis of labelled and unlabelled substances . . . . . . . . . . . . . . . . . . . . 172.2 Validation of synthetic substances . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 172.3 Physicochemical properties . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 172.4 Transformation/Decomposition . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 192.4.1 Microbial . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 192.4.2 Photolytic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20

3. Analytical methods for brominatedflame retardants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 224. Toxicology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 244.1 Uptake, distribution, metabolism, excretion . . . . . . . . . . . . . . . . . . . . . . . 244.1.1 Mammals. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 244.1.1.1 Lower brominated PBDEs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 244.1.1.2 DeBDE . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 264.1.1.3 TBBPA. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 264.1.1.4 HBCD. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 274.1.2 Fish and shellfish. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 284.1.2.1 Lower brominated PBDEs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 284.1.2.2 DeBDE . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 294.1.2.3 TBBPA. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 304.1.2.4 HBCD. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 304.2 Toxicity/effects. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 314.2.1 PBDEs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 314.2.1.1 In vitro . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 314.2.1.2 In mammals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 344.2.1.3 In fish . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 384.2.2 TBBPA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 39

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4.2.2.1 In vitro . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 394.2.2.2 In mammals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 394.2.2.3 In fish . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 404.2.3 HBCD . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 404.2.3.1 In vitro . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 404.2.3.2 In mammals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 404.2.3.3 In fish . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 40

5. Environmental concentrations . . . . . . . . . . . . . . . . . . . . . . . . 415.1 Abiotic samples . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 415.1.1 Air . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 415.1.2 Water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 445.1.3 Sewage sludge . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 445.1.4 Sediment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 455.2 Biological samples . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 525.2.1 Terrestrial ecosystem. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 525.2.1.1 Birds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 525.2.1.2 Mammals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 525.2.1.3 Humans. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 525.2.2 Freshwater ecosystem . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 555.2.2.1 Fish . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 555.2.2.2 Birds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 585.2.3 Marine ecosystem . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 585.2.3.1 Fish and shellfish . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 585.2.3.2 Birds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 595.2.3.3 Mammals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 605.3 Bioaccumulation, biomagnification . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 62

6. Trends . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 706.1 Spatial trends . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 706.2 Temporal trends . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 716.2.1 Baltic Sea sediment core . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 716.2.2 Pike from Lake Bolmen . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 726.2.3 Roach from Lake Krankesjön . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 736.2.4 Guillemot eggs from St. Karlsön. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 746.2.5 Human breast milk . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 76

7. Summary and conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 788. References. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 81

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Abbreviations

AHH Aryl hydrocarbon hydroxylase

CP Chlorinated paraffins

BDE Brominated diphenyl ether

DeBB Decabrominated biphenyl

DeBDE Decabrominated diphenyl ether

DiBDE Dibrominated diphenyl ether

EROD Ethoxy resorufin-O-deethylase

HBCD Hexabromocyclododecane

HxBB Hexabrominated biphenyl

HxBDE Hexabrominated diphenyl ether

HpBDE Heptabrominated diphenyl ether

IUPAC International Union of Pure and Applied Chemistry

MROD Methoxy resorufin-O-deethylase

OcBDE Octabrominated diphenyl ether

PBBs Polybrominated biphenyls

PBDEs Polybrominated diphenyl ethers

PCBs Polychlorinated biphenyls

PCDDs Polychlorinated dibenzo-p-dioxins

PCDFs Polychlorinated dibenzofurans

PeBDE Pentabrominated diphenyl ether

PROD Pentoxy resorufin-O-deethylase

TBBPA Tetrabromobisphenol A

TeBDE Tetrabrominated diphenyl ether

TrBDE Tribrominated diphenyl ether

T4 Thyroxine or 3,3',5,5'-tetraiodo-L-thyronine

T3 3,3',5-triiodothyronine

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Summary in Swedish

Flamskyddsmedel används i plast, gummi och textilier för att förhindrabränder. Direkta källor till utsläpp av bromerade flamskyddsmedel somupptäckts i Sverige är knutna till plast- och textilindustrier och förhöjdahalter har påvisats i sediment och fisk nedströms sådana anläggningar. Stu-dier av damm och inomhusluft visar att dessa substanser även finns i ar-betsmiljön, möjligen genom läckage från plastdetaljer i datorer och annanelektronisk utrustning. Polybromerade difenyletrar (PBDE) finns i dagöverallt i Sverige och till och med i Arktis. Förekomst i utomhusluft indi-kerar att dessa kan spridas via långväga transport. Bromerade flamskydds-medel har även hittats i vatten, sediment och rötslam.

I Sverige är halterna i landlevande däggdjur och fåglar låga, men är hög-re i djur från akvatisk miljö. Relativt höga halter har påvisats i däggdjur ochfisk från Östersjön men också från Nordsjön. Betydligt lägre halter har på-visats i norra Sverige och Arktis. PBDEs förekomstmönster liknar alltsådet som vi sett hos PCB och DDT.

De högsta halterna av PBDE i Sverige har påvisats i sediment (ng/g torr-vikt) och fisk (µg/g fettvikt) längs Viskan i anslutning till textilindustrier.Även DeBDE and HBCD har påvisats i sediment och HBCD i fisk frånViskan. Höga halter TeBDE, PeBDE, OcBDE och DeBDE har hittats i se-diment från flera brittiska floder där utsläpp från industrier som tillverkareller använder PBDE pågår (upp till µg/g torrvikt). Fisk i de påverkadeflodmynningarna visar också höga halter av TeBDE, PeBDE och OcBDE(upp till µg/g fettvikt).

Lägre bromerade PBDE, OcBDE och HBCD är biotillgängliga och åter-finns i fisk. Upptaget via mag-tarmkanalen i råttor, möss och fisk är högtför lägre bromerade PBDE. Musungars upptag av dessa substanser frånbröstmjölk är också högt. Mönstren skiljer sig dock mellan olika arter vil-ket delvis kan bero på skillnader i metabolism. PBDE kan metaboliseras tillhydroxylerade PBDE. Upptaget av HBCD från mag-tarmkanalen är högthos råttor. DeBDE är biotillgänglig även om upptaget är långsamt i fiskmen det verkar som den del som tas upp debromeras i fisken till lägre bro-merade PBDE.

2,2',4,4'-TeBDE (BDE-47), 2,2',4,4',5-PeBDE (BDE-99) och 2,2',4,4',6-PeBDE (BDE-100) biomagnificeras i lax, fiskätande fåglar och däggdjur.

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BDE-47, -99 och -100 har hittats i blod från människa och lax och PBDEfinns även i människors fettvävnad och bröstmjölk. Högre halter BDE-47,hexaBDE (BDE-153 och -154), heptaBDE (BDE-183) och DeBDE har hit-tats i personer som arbetar med återvinning och destruktion av elektroniskutrustning jämfört med människor som arbetar framför datorer eller somlokalvårdare. Flera hydroxylerade och metoxylerade PBDE (Te- och PeB-DE) har hittats i lax, strömming, vikare och gråsäl från Östersjön. Metox-ylerade PBDE har också hittats i gädda från en sötvattensjö. Ursprunget tilldessa ämnen är ej känt.

Flera lägre bromerade PBDE, inklusive BDE-47, -99 och -100, verkarantingen aktiverande eller hämmande via Ah-receptorn (dioxinreceptorn).Bromkal 70-5DE, en PeBDE teknisk produkt, inducerar Ah-receptorme-dierade leverenzymer, t ex EROD, in vitro och i råttor och regnbåge in vi-vo. BDE-47 och -99 har däremot visats sänka EROD-aktiviteten i regn-bågslever. Hydroxylerade PBDE och TBBPA är potenta agonister förtranstyretin (TTR), ett plasmaprotein som transporterar sköldkörtelhormo-ner i blodet. Bromerade analoger till sköldkörtelhormonerna tyroxin (T4)och triiodotyronin (T3) binder också till dessas receptorer. Råttor och mössbehandlade med Bromkal 70-5DE eller BDE-47 visar minskade tyroxin-halter och ändringar i immunsystemet. Möss matade med BDE-47 ellerBDE-99, 10 dagar efter födsel, visar permanenta förändringar i spontantmotoriskt beteende som förvärras med åldern. Liknande exponering tillBDE-99 orsakade även effekter på inlärning och minne när djuren blevvuxna.

Vissa bromerade flamskyddsmedel kan alltså inducera eller nedregleraleverenzymproduktion, påverka tyroideahormonsystemet, vara immuno-toxiska samt neurotoxiska när exponering sker vid en känslig tidpunkt förhjärntillväxt. Mycket litet är känt om TBBPA utöver att det in vitro kanbinda till de plasmaproteiner som transporterar sköldkörtelhormoner i blo-det. Däremot tycks det inte kunna binda till samma protein in vivo. Nästaninget alls är känt om HBCD's effekter.

Studier över tid visar ökade halter av PBDE och HBCD i miljön sedan1970-talet. Tidstrenden i sillgrissla visar att nivåerna av BDE-47, -99 och-100 i Östersjön har börjat sjunka efter mitten av 1980-talet medan tids-trenden i gädda från Bolmen indikerar att halterna inte har förändrats sedanbörjan av 1980-talet. I bröstmjölk däremot har halterna ökat exponentielltmed en dubblering var femte år. Detta kan indikera att exponeringssitua-

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tionen är olika i Östersjön, sötvattensekosystem och för människor. Män-niskor exponeras inte bara via maten utan även från den miljö man lever i,i hemmet och på arbetsplatsen, vilket kan påverka den samlade expo-neringen.

PBDE, TBBPA och HBCD finns i miljön. De tas upp av levande orga-nismer och åtminstone de lägre bromerade PBDE biomagnificeras. TBB-PA och PBDE och/eller deras metaboliter är biologiskt aktiva. Halterna avPBDE verkar öka, och trenden i människor indikerar att ökningen kan varasnabb. Om inte denna trend bryts kommer den att leda till att belastningeni vilda djur och i människor når nivåer som orsakar effekter. Vår kunskapom dessa substanser, deras källor, toxicitet och beteende i miljön är än sålänge mycket begränsad, vilket gör riskbedömningen osäker. SärskiltPBDE är ett miljöproblem på grund av deras persistens, lipofilicitet ochbioackumulerbarhet. Kunskapen om TBBPA och HBCD är så begränsadatt det idag är omöjligt att ens uppskatta vilken risk de utgör. Tillsammansindikerar dessa resultat att vi kan stå inför ett nytt ”PCB problem”.

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1. Introduction

1.1 What are flame retardants?Flame retardants are substances used in plastics, textiles, electronic circu-itry and other materials to prevent fires. Some of the technical flame retard-ant products contain brominated organic compounds including polybromi-nated biphenyls (PBBs), polybrominated diphenyl ethers (PBDEs), tetra-bromobisphenol A (TBBPA) and hexabromocyclododecane (HBCD). Thestructures for these are shown in Figure 1. Many of these substances arepersistent and lipophilic and have been shown to bioaccumulate.

Some brominated flame retardants are additives mixed into polymersand are not chemically bound to the plastic or textiles and are therefore ableto leak out of products and into the environment. Additives include PBBs,

Figure 1. The chemical structures of a) polybrominated diphenyl ethers, b)hexabromocyclododecane (HBCD), c) tetrabromobisphenol A (TBBPA), and d)polybrominated biphenyls.

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PBDEs and HBCD. Others, such as TBBPA, are reactive and are bound tothe material chemically. However, some of the reactive flame retardantmay not have polymerized and can thus leak out of the product.

Brominated flame retardants as such are not produced in Sweden. Forthe most part, these substances are imported into Sweden in finished prod-ucts and goods or in materials used to make products and goods. The cur-rent environmental levels may partly be the result of leaching from firstgeneration flame-retarded products dumped in landfills over a decade ago.Knowledge about these substances is very limited and hinders environ-mental authorities from carrying out adequate risk assessments.

The major companies producing brominated flame retardants are EthylCorporation (USA), Great Lakes Chemical Corporation (USA and UK),Dead Sea Bromine (Israel) and Eurobrom (Netherlands). Other companiesinclude Riedel de Haen (Hoechst Group), Ceca (ATOCHEM, France), Po-tasse et Produit Chimiques (Rhone Poulenc Group), Warwick Chemicals(UK), Albemarle S.A. (Belgium) as well as Nippo, Tosoh and Matsunagaall from Japan (KEMI, 1994a; WHO/IPCS, 1994b). The total world pro-duction of all brominated flame retardants is approximately 150 000 metrictons/year. Forty percent is distributed to North America, 30% to the FarEast and 25% to Europe (KEMI, 1994a).

1.2 Which substances are coveredin this report?PBBs have not been prioritized in this report for several reasons. Due to amajor PBB-poisoning accident in the USA in 1973, substantial knowledgeof the toxicology of some PBBs is now available (KEMI, 1990; WHO/IPCS, 1994a). For a recent review of PBBs see de Boer et al. (2000). Also,the environmental levels of PBBs are low in Sweden. Thus, the knowledgebase for risk assessment of PBBs in Sweden is more satisfactory. The ma-jor emphasis of this report is therefore on PBDEs, and to some extent onTBBPA and HBCD, as knowledge about these substances is incomplete orin some cases, almost non-existent. Other brominated organic substancesare also of concern, such as polybrominated dibenzo-p-dioxins and furans.PBDD/Fs are formed, for example, when plastics containing brominatedflame retardants are heated (welding of mats, melting of polymers). How-ever, coverage of PBDD/Fs is outside the scope of this report. Interested

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readers are referred to recent reports on PBDEs and PBDD/Fs by WHO/ICPS for more information (WHO/ICPS, 1994b; 1997; 1998).

1.2.1 POLYBROMINATED DIPHENYL ETHERS (PBDES)There are theoretically 209 PBDE congeners. Technical PBDE productsare produced by brominating diphenyl ether in the presence of a catalyst.The major technical products contain mainly pentaBDEs, octaBDEs or de-caBDE, but contain other PBDEs as well. The general compositions of thetechnical products are given in Table 1. The individual PBDE congenersare numbered according to the IUPAC system used for numbering PCBsbased on the position of the halogen atoms on the rings.

Table 1. The general compositions of PBDE-based flame retardants given in percent ofBDE congeners present (WHO/IPCS, 1994b).

PBDEs are mainly imported to Sweden incorporated in plastics used inproduction or in finished products such as TVs, computers, electricalequipment, household appliances, cables, furniture, textiles, etc. Annualworldwide production of penta-, octa- and decaBDE technical products in1990 was estimated to be 4 000, 6 000 and 30 000 metric tons respectively(Arias, 1992). Consumption of PBDEs for 1999 in the European Unionwas estimated to be 150 metric tons penta-, 400 metric tons octa- and 7 000metric tons decaBDE technical products (De Poortere, 2000). Accumulat-ed amounts in electric products still in use in 1994 in the Nordic countrieswere estimated to be 250 metric tons of PeBDE and 5500 metric tons ofOcBDE and DeBDE combined (Hedelmalm et al., 1995). The estimatedamounts supplied to Sweden alone in 1991 were 17 metric tons of PeBDEand 626 metric tons of OcBDE and DeBDE combined (Hedelmalm et al.,1995). The combined total annual amounts of penta-, octa- and decaBDEtechnical products recently imported to Sweden are given in Table 2. For1993, approximately 20 metric tons of PBDE were imported to Sweden

Congener percent

TechnicalProduct

Tetra-BDEs

Penta-BDEs

Hexa-BDEs

Hepta-BDEs

Octa-BDEs

Nona-BDEs

Deca-BDE

PeBDE 24–38 50–60 4–8

OcBDE 10–12 44 31–35 10–11 <1

DeBDE <3 97–98

14

and the same amount was imported in plastics (KEMI, 1995). The amountof PBDE imported in ready-made products the same year was estimated bythe Swedish National Chemicals Inspectorate to be approximately 400metric tons (KEMI, 1995).

Based on this information, technical products containing DeBDE are es-timated to be used in larger quantities than those containing TeBDEs andPeBDEs.

Table 2. The import of brominated flame retardants to Sweden in metric tons (KEMI, 1999).Due to confidentiality regulations, only the sum of penta-, octa- and decabrominated prod-ucts as well as the sum of TBBPA and its carbonate oligomer are available.

*Only decabrominated products were imported.

1.2.1.1 Lower brominated PBDEs (TeBDEs, PeBDEs, HxBDEs)PBDEs are lipophilic and have some structural similarities to PCB andPCDD/F. Bromkal 70-5DE is one of several PeBDE technical productsthat contains lower brominated PBDEs such as 2,2',4,4'-TeBDE (BDE-47),2,2',4,4',5-PeBDE (BDE-99) (Sundström and Hutzinger, 1976), 2,2',4,4',6-PeBDE (BDE-100), as well as two newly identified triBDEs, one addition-al tetraBDE, one additional pentaBDE, three hexaBDEs and one hep-taBDE (Sjödin et al., 1998). PBDEs were first discovered in Sweden in fishsamples taken downstream from several textile industries on the RiverViskan (Andersson and Blomkvist, 1981). In some countries, industries arevoluntarily replacing the lower brominated PBDEs with other flame retard-ants.

1.2.1.2 Higher brominated PBDEs (Hepta-NonaBDEs, DecaBDE)OcBDE technical products are used in acrylonitrile butadiene styrene,polycarbonate and thermosets. DeBDE products are used in most types ofsynthetic materials including textiles and polyester used for printed circuitboards (OECD, 1994). DeBDE is the most widely used technical productof the PBDEs. DeBDE has been shown to photolytically degrade in labo-ratory experiments (Watanabe and Tatsukawa, 1987; U. Örn, Stockholm

Year 1993 1994 1995 1996 1997 1998

PBDE 22 90 20 79 123 41*

HBCD 50 80 90 83 125 89

TBBPA 488 450 258 291 303 269

15

University, personal communication; Sellström et al., 1998a) and formlower brominated PBDE. If this process is significant, it could lead to theformation of TeBDEs, PeBDEs and HxBDEs, which are known to accu-mulate in living organisms.

1.2.2 TETRABROMOBISPHENOL A (TBBPA)TBBPA is covalently bound to plastic and is used in electronic circuitboards. Annual worldwide production has been estimated to be 50 000metric tons/year (KEMI, 1994a). Estimated accumulated amounts of TBB-PA in products in the Nordic countries in 1994 were 4000 metric tons(Hedelmalm et al., 1995). The estimated amount supplied to Sweden alonein products in 1991 was 334 metric tons. More recent information on theimport of TBBPA to Sweden is given in Table 2. One moiety of TBBPA’smolecular structure is similar to that of the thyroid hormone thyroxine, ex-cept that the iodine atoms have been replaced by bromines (see Figure 1c,Figure 2). The dimethylated derivative of TBBPA (MeTA) may have someuse as a flame retardant, but may also be the result of methylation of TBB-PA in sediment (Watanabe et al., 1983).

1.2.3 HEXABROMOCYCLODODECANE (HBCD)HBCD is produced by bromination of cyclododecane in a batch process.HBCD has been used for about 20 years. Approximately 11 000 metrictons are incorporated in articles made of polystyrene and in textile back-coating for the EU market. It is used in foams and expanded polystyrene.End products include upholstered furniture, interior textiles, automobile

Figure 2. The chemical structures of the thyroid hormones 3,3',5,5'-tetraiodo-L-thyronine (thyroxine or T4) and its 5'-deiodinated congener 3,3',5-triiodo-thyronine (T3).

16

interior textiles, car cushions, insulation blocks in trucks and caravans aswell as in building materials such as house walls, cellars, roofs and parkingdecks, against frost heaving in roads and railway embankments, packagingmaterial, video cassette recorder housing and electric equipment (Marga-reta Palmquist, National Swedish Chemicals Inspectorate, personal com-munication). The amounts of HBCD imported to Sweden for the last fewyears are given in Table 2.

1.2.4 POLYBROMINATED BIPHENYLS (PBBS)In 1973, a commercial flame retardant containing PBBs was accidentlymixed into feed for dairy cattle, livestock and poultry in the state of Mich-igan, USA (KEMI, 1990; WHO/IPCS, 1994a). The feed was used widely,leading to widespread PBB-contamination of milk, meat and eggs and poi-soning in animals. Over 9 million people were exposed to PBBs from food.Because of this widespread exposure, research was funded to better under-stand the toxicology of PBBs, and poisoned animals and exposed humanshave been studied as well. The effects of PBBs were found to be essentiallythe same as those seen for PCBs.

Technical hexabrominated biphenyl (HxBB) is banned in North Ameri-ca and in Europe. Technical decabrominated biphenyl (DeBB) is still pro-duced in Europe, but production will be discontinued after 2000 (De Poor-tere, 2000).

17

2. Brominated flame retardant chemistry

2.1 Synthesis of labelled and unlabelled substancesConsiderable work has been carried out to synthesize pure PBDE conge-ners for use as standards, for toxicology studies and in the identification ofunknown substances in environmental samples. Within several researchprograms, a number of both unlabelled and radiolabelled PBDEs have beensynthesized for these purposes (Örn et al., 1996; Hu, 1996; Jakobsson et al.1996; Örn, 1997; Örn et al., 1998; Marsh et al., 1999). This work is sum-marized in Bergman, (1999) and in Table 3.

2.2 Validation of synthetic substancesUsing synthesized single congeners of some PBDEs, it has been possibleto better quantify the components of the technical product Bromkal 70-5DE and to identify the previously unknown Pe1BDE (2,2',4,4',6-PeBDE)in this product. According to previously published results, Bromkal 70-5DE consisted of 41% 2,2',4,4'-TeBDE, 45% 2,2',4,4',5-PeBDE and 7%2,2',4,4',6-PeBDE (Sundström and Hutzinger, 1976). Using the new stand-ards, it has been found that the composition is actually 37%, 35% and 6.8%for these three PBDEs, respectively (Sjödin et al., 1998). Bromkal 70-5DEalso contains 1.6% of 2,2',3,4,4'-PeBDE (BDE-85) and 3.9, 2.5 and 0.41%of HxBDEs -153, -154 (2,2',4,4',5,6'-HxBDE) and -138, respectively. Anumber of other standards that have been produced have also been ana-lyzed in order to use them in identifying and quantifying more PBDEs inenvironmental samples on a congener-specific basis.

2.3 Physicochemical propertiesThe PBDEs have low vapor pressures and are very lipophilic with LogKows (octanol-water partitioning coefficients) ranging from 5.9-6.2 forTeBDEs, 6.5–7.0 for PeBDEs, 8.4-8.9 for OcBDEs and 10 for DeBDE(Watanabe and Tatsukawa, 1990). TBBPA has a Log Kow of 4.5 (WHO/

18

IPCS, 1995). The dimetylated derivative of TBBPA (MeTA) has a Log Kow

of 6.4, making it more lipophilic than the parent compound (Watanabe andTatsukawa, 1990). The Log Kow for HBCD is 5.8 (IUCLID, 1996).

Table 3. Names and congener numbers of PBDEs synthesized.

Name Congenernumber

Reference

2-bromodiphenyl ether BDE-1 Jakobsson et al., 1996; Marsh et al., 1999

3-bromodiphenyl ether BDE-2 Jakobsson et al., 1996; Marsh et al., 1999

4-bromodiphenyl ether BDE-3 Jakobsson et al., 1996; Marsh et al., 1999

2,4-DiBDE BDE-7 Jakobsson et al., 1996; Marsh et al., 1999

2,4'-DiBDE BDE-8 Jakobsson et al., 1996; Marsh et al., 1999

2,6-DiBDE BDE-10 Jakobsson et al., 1996; Marsh et al., 1999

3,4-DiBDE BDE-12 Jakobsson et al., 1996

3,4'-DiBDE BDE-13 Jakobsson et al., 1996; Marsh et al., 1999

4,4'-DiBDE BDE-15 Jakobsson et al., 1996; Marsh et al., 1999

2,2',4-TrBDE BDE-17 Jakobsson et al., 1996; Marsh et al., 1999

2,3',4-TrBDE BDE-25 Jakobsson et al., 1996; Marsh et al., 1999

2,4,4'-TrBDE BDE-28 Jakobsson et al., 1996; Marsh et al., 1999

2,4,6-TrBDE BDE-30 Jakobsson et al., 1996; Marsh et al., 1999

2,4',6-TrBDE BDE-32 Jakobsson et al., 1996; Marsh et al., 1999

2',3,4-TrBDE BDE-33 Marsh et al., 1999

3,3',4-TrBDE BDE-35 Jakobsson et al., 1996

3,4,4'-TrBDE BDE-37 Jakobsson et al., 1996

2,2',4,4'-TeBDE BDE-47* Örn et al., 1996; Jakobsson et al., 1996; Örn,1997; Marsh et al., 1999

2,2',4,5'-TeBDE BDE-49 Marsh et al., 1999

2,2',4,6'-TeBDE BDE-51 Jakobsson et al., 1996; Marsh et al., 1999

2,3',4,4'-TeBDE BDE-66 Jakobsson et al., 1996

2,3',4',6-TeBDE BDE-71 Jakobsson et al., 1996

2,4,4',6-TeBDE BDE-75 Jakobsson et al., 1996; Marsh et al., 1999

3,3',4,4'-TeBDE BDE-77 Marsh et al., 1999

2,2',3,4,4'-PeBDE BDE-85* Örn et al., 1996; Örn,, 1997

2,2',4,4',5-PeBDE BDE-99* Örn et al., 1996; Örn, 1997

2,2',4,4',6-PeBDE BDE-100 Jakobsson et al., 1996; Marsh et al., 1999

19

*Radiolabelled congeners using 14C were also synthesized.

PBDEs are persistent, have low water solubility, high binding affinity toparticles and a tendency to accumulate in sediments. HBCD also has lowwater solubility (IUCLID, 1996), and probably also has an affinity for par-ticles and sediments.

2.4 Transformation/Decomposition2.4.1 MICROBIAL

A study of anaerobic microorganisms’ ability to break down DeBDE tolower brominated PBDE in sediment was carried out during 1994. DeBDEwas applied to anaerobic sediment which was then inoculated with micro-organisms enriched from a PBDE-contaminated sediment. The sedimentwas then divided into smaller samples and allowed to gently shake. Sam-ples were analyzed at different time points but no breakdown of DeBDEwas seen during the experimental time of four months. The experiment wasextended by letting one aliquot of sediment continue incubation. Subsam-ples were analyzed at several time points but no breakdown could be seenafter an incubation of 2 years (unpublished results, Ulla Sellström; de Wit,1995; 1997).

Studies of TBBPA in sediments have shown that a dimethylated deriva-tive is also found (Sellström and Jansson, 1995). The origin of this meth-ylated TBBPA is not completely understood, however it probably has

2,3,4,5,6-PeBDE BDE-116 Jakobsson et al., 1996; Marsh et al., 1999

2,3',4,4',6-PeBDE BDE-119 Jakobsson et al., 1996

2,2',3,3',4,4'-HxBDE BDE-128 Örn et al., 1996; Örn, 1997

2,2',3,4,4',5'-HxBDE BDE-138 Örn et al., 1996; Örn, 1997

2,2',3,4,4',6'-HxBDE BDE-140 Marsh et al., 1999

2,2',4,4',5,5'-HxBDE BDE-153 Örn, et al., 1996; Örn, 1997

2,2',4,4',5,6'-HxBDE BDE-154 Marsh et al., 1999

2,3,4,4',5,6-HxBDE BDE-166 Jakobsson et al., 1996; Marsh et al., 1999

2,2',3,4,4',5,6-HpBDE BDE-181 Jakobsson et al., 1996; Marsh et al., 1999

2,3,3',4,4',5,6-HpBDE BDE-190 Jakobsson et al., 1996

Name Congenernumber

Reference

20

some use as a flame retardant (WHO/IPCS, 1997) but may also be due tomethylation of TBBPA by microorganisms in the sediment (Watanabe etal., 1983).

2.4.2 PHOTOLYTIC

Previous studies indicate that DeBDE is debrominated by UV light andsunlight to lower brominated PBDE (to triBDE with UV light and totetraBDE with sunlight) but it is not known if this happens in the environ-ment (Norris et al., 1973, 1975a; Watanabe and Tatsukawa, 1987; UlrikaÖrn, Department of Environmental Chemistry, Stockholm University,pers. comm.).

Recent laboratory studies of the photolytic breakdown of DeBDEshowed that DeBDE in toluene and applied to silica gel is successively de-brominated by UV light to lower brominated PBDE (down to TeBDEs)and that this occurs very rapidly (Sellström et al., 1998a). The half-life intoluene was less than 15 minutes. DeBDE-treated sand, soil and sedimentwere exposed to UV light in the laboratory or to sunlight outdoors for dif-ferent time periods and then analyzed. Results from the sand series exposedto both UV and sunlight show that DeBDE is photolytically debrominatedin the same manner as in solution and on silica gel, but that the debromina-tion time course proceeds more slowly for both UV and sunlight exposure(half-life of 12 and 37 hours, respectively). Similar results were obtainedfor sediment, with a half-life for DeBDE of 53 hours for UV-exposure and81 hours for sunlight exposure and with TeBDEs appearing at the longestexposure time (244 hours). Anomalous results were obtained for the soilsamples exposed to sunlight, making interpretation difficult. However, theresults for soil samples treated with DeBDE and exposed to UV-light in thelaboratory indicated a half-life of 185 hours and the debromination processwas the same as that seen for all the other matrices tested.

TBBPA is also photolytically decomposed when exposed to UV light,both in the absence and presence of hydroxyl radicals (Eriksson and Jakobs-son, 1998). The main breakdown product is 2,4,6-tribromophenol (Figure3). A number of other decomposition products are also found and some ofthese have been tentatively identified as di- and tri-bromobisphenol A, di-

21

bromophenol, 2,6-dibromo-4-(bromoisopropylene)phenol, 2,6-dibromo-4-(dibromoisopropylene)phenol and 2,6-dibromo-1,4-hydroxybenzene.

0

0,5

1

1,5

2

2,5

3

0 20 40 60 80 100

time(h)

Rel

. int

ensi

ty T

BB

PA

0

0,01

0,02

0,03

0,04

Rel

. i n

tens

ity T

BP

TBBPA

TBP

Expon. (TBBPA)

Figure 3. Decomposition of TBBPA and formation of tribromophenol (TBP)(Eriksson and Jakobsson, 1998).

22

3. Analytical methods for brominated flame retardantsIn most analytical methods for brominated organic compounds, the sampleis first extracted with an organic solvent. Lipids can be removed using sul-furic acid treatment or gel permeation chromatography methods. In somecases the extract needs to go through a further clean-up step using someform of column chromatography to remove interfering substances. The pu-rified extract is then analyzed with gas chromatography using electron cap-ture detection (ECD), mass spectrometry using the negative ions formed atchemical ionization (MS-ECNI) (Jansson et al., 1987; Jansson et al. 1991;Sellström et al., 1993a; 1998a; Nylund et al. 1992) or high resolution gaschromatography-masspectrometry (HRGC-MS)(Meironyté et al., 1999).The higher brominated PBDEs have longer retention times and are oftenanalyzed using a shorter GC column (Sellström, 1996; Sellström et al.,1998a). For a more detailed review of analytical methods, see de Boer etal. (2000).

TBBPA is separated from the neutral components by treatment with abasic water solution. The pH of this water solution is then adjusted to beacidic and TBBPA is extracted with an organic solvent. Before analysis,TBBPA is derivatized to its diacetylated derivative. Dimethylated TBBPA(MeTA) is analyzed with the PBDEs (Sellström and Jansson, 1995).

Recovery experiments have been performed for TeBDE (BDE-47),PeBDE (BDEs -99, -100) and DeBDE (BDE-209) in fish muscle and forBDEs -47, -99, -100, -209 and HBCD in sediment (Sellström et al. 1998b).The results showed recoveries of 111–114% for fish muscle and 106-140%for sediment (Sellström et al., 1998b).

It is difficult to quantify HxBBs because of interference by HxBDEs inthe analysis. An HPLC method was therefore developed to separate PBBsfrom PBDEs using an HPLC amino column. ”Clean” fish samples werespiked with known amounts of several brominated flame retardants (tetra-hexaBBs, methylated TBBPA, TeBDE, PeBDE, HxBDE, DeBDE, HB-CD), extracted and the extracts fractionated using the new separation meth-od. To further validate the method, a few fish and sediment samples fromRiver Viskan were extracted and the extracts separated using the new

23

method. The method was successful in separating the PBBs from PBDEsand the levels of PBBs were found to be very low in these samples (Sell-ström et al., 1994).

Hydroxylated PBDE are separated from the neutral components bytreatment with a basic water solution and then derivatized with diazometh-ane. Methoxy-PBDE are found in the neutral fraction. The analysis is car-ried out using gas chromatography-mass spectrometry or GC-ECD (Asp-lund et al., 1997; 1999a; Haglund et al., 1997).

24

4. ToxicologyThorough reviews of previous studies of the toxicology of PBBs, PBDEsand TBBPA can be found in the relevant WHO/IPCS reports (1994a;1994b; 1995) and in Darnerud et al. (1998) for PBDEs.

4.1. Uptake, distribution,metabolism, excretion 4.1.1 MAMMALS

4.1.1.1 LOWER BROMINATED PBDES

The distribution of 14C-labelled 2,2',4,4'-tetrabromodiphenyl ether (BDE-47), 2,2',3,4,4'-pentabromodiphenyl ether (BDE-85), and 2,2',4,4',5-penta-bromodiphenyl ether (BDE-99) was studied in C57BL mice using whole-body autoradiography, and the latter two congeners were also used in milktransfer studies during the neonatal period, by use of liquid scintillationtechnique (Darnerud and Risberg, 1998).

The autoradiograph findings in this study showed that the uptake fromthe gastrointestinal tract was effective. Organs and tissues with high radio-activity concentrations were the fat depots, liver, adrenal and ovary, lungand (shortly after administration) the brain. At longer post-injection time,the apparent concentration in most tissues was considerably lower, and ra-dioactivity was still present in white and brown fat depots. Studies in preg-nant mice showed a low fetal uptake of the compounds. No significant dif-ference in distribution between the three studied congeners was observed.Breast milk transport of the pentaBDE congeners was substantial and 30–40% of the administered single dose of radioactivity was found in the suck-ling offspring litter after four days. At the same time point, the plasma lev-els in the neonates were more than two times that of the mothers, althoughthe absolute levels were low. The concentration in tissues and milk was inabout the same range as that earlier observed for PCB congeners of a sim-ilar degree of halogenation, when administered in equimolar doses.

Uptake, distribution, metabolism and excretion of BDE-47 has beenstudied in rats and mice dosed orally with 14C-labelled BDE-47 (Klasson-Wehler et al., 1996; Örn, 1997; Örn and Klasson-Wehler, 1998). In rats,

25

14% was excreted in feces and less than 0.5% in urine during 5 days. Ofthe amount excreted, 79% was the parent compound and 21% correspond-ed to metabolites. Treated rats retained 86% of the administered dose after5 days and the highest concentrations were found in adipose tissue, wherethe 14C corresponded to the parent BDE-47. All tissues analyzed containedparent BDE-47. Liver also contained low concentrations of five hydroxy-lated metabolites. Two of these were found in plasma (Örn och Klasson-Wehler, 1998).

In mice, 20% of the dose was excreted in feces and 33% in urine. Of theamount excreted, 15% was parent compound and 85% corresponded tometabolites. Treated mice retained 47% of the administered dose after 5days and similar concentrations were found in both adipose tissue and liv-er. Covalently bound metabolites were found in liver (12% bound to mac-romolecules, 16% to lipids), in the lungs (28%) and kidneys (4%). Lowconcentrations of five hydroxylated metabolites were also observed in theliver. Three of these were seen in plasma as well (Örn och Klasson-Wehler,1998). The conclusions drawn from this study were that BDE-47 was ab-sorbed well by both the rat and mouse, but the rate of metabolism and ex-cretion varied considerably.

Tissue disposition, metabolism and excretion of BDE-99 has been stud-ied in bile-cannulated and uncannulated male rats dosed orally with 14C-labelled BDE-99 (Hakk et al., 1999; Larsen et al., 1999). Feces was themajor route of elimination in both groups (43% of the administered dose inuncannulated and 86% in cannulated rats after 72 h). Of the amount excret-ed in feces, more than 90% was parent BDE-99. Cumulative excretion ofmetabolites in both groups of rats was less than 1% of the administereddose into urine and 3.7% into bile. In the uncannulated rats, 6.3% of the 14Cin urine was protein bound and was associated with alpha-2-microglobulin.In cannulated rats, 28–47% of the 14C in bile was bound to an unidentifiedprotein of 79 kDa. In the uncannulated rats, 39% of the administered dosewas retained and the highest concentrations were found in adipose tissue,skin and adrenals. Covalently bound 14C was observed in feces indicatingthe formation of reactive metabolic intermediates. Two monomethoxypentabromodiphenyl ether metabolites and two de-brominated monometh-oxy tetrabromodiphenyl ether metabolites were indicated in feces. Twomonohydroxy pentabromodiphenyl ether metabolites, two dihydroxy

26

pentabromodiphenyl ether metabolites and possibly two thio-substitutedpentabromodiphenyl ether metabolites were characterized in bile.

In a recent study, 17 PBDE congeners (BDEs -15, -28, -30, -32, -47, -51,-71, -75, -77, -85, -99, -100, -119, -138, -153, -166 and -190) were incubat-ed individually with rat hepatic microsomes from rats treated with beta-naphthaflavone, phenobarbital or clofibrate (Meerts et al., 1998b). Theoriginal congeners and the metabolites formed were then tested for theirability to compete with the thyroid hormone thyroxine (T4) for binding tohuman transthyretin in vitro. A number of hydroxylated PCBs are structur-al analogues of T4 and are known to compete with T4 in this system. Thestructural requirements for binding to transthyretin are hydroxy-substitu-tion in para- or meta-positions of one or both phenyl rings with adjacenthalogen substitution (Lans, 1995). X-ray diffraction studies have shownthat several phenolic organohalogen compounds bind with the hydroxygroup in the central channel of the transthyretin molecule.

Results showed no competition with the parent compounds, but consid-erable potency for several of the metabolites, indicating the metabolism ofPBDE to hydroxylated PBDE (see 4.2.1.1 for results). No binding compe-tition was seen for several of the higher brominated PBDE such as BDEs -138, -153, -166 and -190 after incubation with the microsomes. This mayindicate that these congeners are not readily metabolized.

4.1.1.2 DeBDEThe uptake of 14C-DeBDE has previously been shown to be low followingoral administration to rats, and between 90% to 99% of the dose was elim-inated in the feces and gut (Norris et al., 1975b; El Dareer et al., 1987). A2-year feeding study of DeBDE in rats confirmed a low accumulation,however, the small portion that reached the adipose tissue, measured as thetotal bromine content, remained unaffected for 90 days of recovery (Norriset al., 1975b; Kociba et al., 1975). When 14C-DeBDE was injected intra-venously, 74% of the dose was found in feces and gut contents after 72hours. Of the excreted material, 63% was metabolites and 37% was the par-ent compound (El Dareer et al., 1987).

4.1.1.3 TBBPATBBPA fed to rats was eliminated primarily in feces (95%) and 1% waseliminated in urine within 3 days (WHO/IPCS, 1995).

27

14C-labelled TBBPA was fed to rats with and without bile duct cannula-tion (Larsen et al., 1998). Bile, urine and feces were collected every 24hours for three days, after which the rats were killed and organs and tissuessampled. Three conjugated metabolites were found in the bile – a diglu-curonide, a monoglucuronide and a glucuronide-sulfate ester. These repre-sented 34, 45 and 21% of the radioactivity excreted in bile during the first24 hours, respectively. In the cannulated rats, 71% of the dose was excretedin the bile within 72 hours. In uncannulated rats, 95% of the 14C-labelledTBBPA was excreted in the feces as TBBPA. However, there was a delayin fecal excretion. This is the result of enterohepatic circulation of TBBPAwhere the biliary metabolites are deconjugated and reabsorbed from thelower intestine, reconjugated and re-excreted in the bile. About 2% of thedose remained in the uncannulated rat after 72 hours and highest levelswere found in the large intestine (1%), small intestine (0.6%), lung (0.2%)and carcass (0.2%).

Uptake and distribution of 14C-labelled TBBPA in pregnant rats afteroral exposure on gestational days 10 to 16 has recently been studied(Meerts et al., 1999). The major portion of radioactivity was excreted infeces (79.8%). Only 0.83% of the total administered dose was found in thetissues of the dams and 0.34% in the fetuses. Highest maternal levels werefound in the carcass (0.37%) and liver (0.26%).

4.1.1.4 HBCDRats fed a single oral dose of 14C-labelled-HBCD rapidly absorbed the sub-stance (Yu and Atallah, 1980). HBCD was readily distributed in the entirebody with highest concentrations found in adipose tissue, followed by liv-er, kidney, lung and gonads. The half-life was two hours. HBCD was rap-idly metabolized and 72% of the dose was eliminated via feces and 16% byurine within 72 hours. Four metabolites were found but no information ontheir structures was given. HBCD absorption followed a two-compartmentopen model system, with the central compartment consisting of blood,muscle, liver, kidney and other non-adipose tissues, and the peripheralcompartment consisting of fat tissue. Elimination from fat was slower thanfor the central compartment.

Rats fed HBCD daily for 5 days showed no urinary excretion of HBCD(Ryuich et al., 1983). Average daily fecal excretion was 29–37% of the ad-ministered amount. HBCD was found to accumulate in adipose tissue in

28

this study. Experiments using a loop of the upper jejunum suggest thatHBCD can be absorbed from the intestine.

4.1.2 FISH AND SHELLFISH

4.1.2.1 LOWER BROMINATED PBDEA study of uptake, accumulation and excretion of BDE-47, -99 and -153from water was carried out in blue mussels (Edulis mytilus) (Gustafsson etal., 1999). Several PCB congeners were included for comparison (tri-hexa-CBs). The uptake clearance rates were found to be approximately ten timeshigher for BDE-47 and -99 than for BDE-153 and the PCB congeners.Depuration rates were similar for all three PBDEs indicating no depend-ence on hydrophobicity, but were correlated to hydrophobicity for thePCBs.

A dietary uptake study of TeBDE, PeBDE and HxBDE has been carriedout in pike (Burreau et al., 1997). Pike were fed with rainbow trout that hadbeen injected with a mixture containing BDE-47, -99 and -153 as well asselected PCB and PCN congeners. The uptake efficiencies of all three PB-DEs were high, with BDE-47 showing the highest uptake (more than 90%of the given dose) from the gastrointestinal tract. Uptake efficiencies forBDE-99 and – 153 were 62% and 40% repectively. The uptake efficiencyof BDE-47 was higher than for the tri- to hexaCBs and the hexa- to octaC-Ns tested. The uptake efficiencies were higher than expected consideringthe size and lipophilicity of the compounds. It was concluded that uptakeof these substances from the gastrointestinal tract may be facilitated bycotransport with lipids and/or proteins through a mediated or even activetransport mechanism.

In a distribution study, several pike were fed rainbow trout previouslyinjected with 5 microcuries of 14C-2,2',4,4'-TeBDE (BDE-47). Pike werekilled after 9, 18, 36 and 65 days and whole-body autoradiography per-formed. The results showed accumulation of the labelled TeBDE in liver,gall blader, kidneys, brain, chorion of the eye, in the perivisceral adiposetissue and along the spinal column (Figure 4) (Burreau and Broman, 1998;Burreau et al., 2000).

Asplund et al. (1997; 1999a) have found hydroxylated and methoxylat-ed PBDE in Baltic salmon blood plasma, and Haglund et al. (1997) foundmethoxylated PBDE in ringed and grey seal, salmon and herring from the

29

Baltic Sea. Kierkegaard et al. (1999) found methoxy-BDE-47 in pike fromLake Bolmen, a freshwater lake. It is not clear what the source of these arebut one possibility may be due to metabolism either in the organisms or bymicroorganisms. Natural production by invertebrates or algae can not beruled out.

4.1.2.2 DeBDEA major argument made by flame retardant producers for the use ofDeBDE is that the molecule is so large that it isn’t bioavailable and there-fore will not be accumulated by living organisms. To determine if this wasthe case, a dietary uptake study of DeBDE in rainbow trout was carried out(Kierkegaard et al., 1994; Kierkegaard et al., 1999a). Some biological ef-fects were also measured. Trout were fed with food containing DeBDE (10mg DeBDE/kg/day) for 0, 16, 49 and 120 days. One group was treated withclean food for 71 days after 49 days of exposure to study elimination.TheDeBDE concentrations in muscle ranged from <0.6 ng/g fresh weight after0 days, to 38 (±14) ng/g fresh weight after 120 days. Corresponding liverconcentrations were <5 and 870 (±219) ng/g fresh weight. A number of or-ganic brominated substances, characterized as hexa- to nonaBDEs, in-creased in concentration with exposure length.

Gall Bladder Muscle Eye

Vertebrae Surrounding Tissue

Kidney Liver Heart

Perivisceral Adipose Tissue

Skin

Figure 4. Tape section radiogram showing the distribution of radiolabelled2,2',4,4'-TeBDE in pike 9 days after dietary exposure (Burreau and Broman,1998).

30

After the 71 day depuration period, DeBDE levels declined, but the lev-els of some of the lower brominated congeners were unaffected during thesame period. This may indicate that DeBDE is metabolized via a reductivedebromination process to the lower brominated diphenyl ethers and/or thatlower brominated PBDE present in the technical product are selectively ab-sorbed. However, no HxBDEs were detected in the technical product andthere was a pronounced shift in increasing peak heights for the first-elutingcongeners in the fish compared to the technical product they were fed. Theestimated uptake of DeBDE was 0.02-0.13% based on DeBDE concentra-tions in muscle and an estimation of the concentrations of the metabolitesformed compared to the total dose administered.

A number of physiological and biochemical variables were investigated.Liver body index (indicating increased liver weight) and plasma lactatelevels increased in fish exposed for 120 days and in the depuration group.The number of lymphocytes was significantly lower after 120 days expo-sure compared to controls. DeBDE did not affect ethoxyresorufin-O-deethylase (EROD), ethoxycoumarin-O-deethylase (ECOD) or transketo-lase activity and no DNA adducts were seen.

4.1.2.3 TBBPATBBPA is rapidly taken up from water via the gills in fish. When fish areplaced in clean water, the compound is rapidly eliminated (WHO/IPCS,1995).

4.1.2.4 HBCDThirty fathead minnows (Pimephales promelas) were exposed to HBCD inwater for 32 days (Veith et al., 1979). Five fish were removed for analysisafter 2, 4, 8, 16, 24 and 32 days of exposure. The bioconcentration factorwas estimated to be 18 100 (log BCF 4.26). Based on these results, HBCDseems to have high bioaccumulating potential.

In a study along the River Viskan, HBCD was found in both sedimentsand in pike (Sellström et al., 1998b), indicating that HBCD is bioavailable.It is probable that the levels found in pike are due to a combination ofuptake via the gills and via food (gastrointestinal tract).

31

4.2 Toxicity/effects As has been stated previously, more detailed reviews of the toxicology, in-cluding the toxicity and effects of PBBs, PBDEs and TBBPA, can be foundin the relevant WHO/IPCS reports (1994a; 1994b; 1995).

4.2.1 PBDES

4.2.1.1 IN VITRO A PeBDE technical product, Bromkal 70-5DE, has weak dioxin-like tox-icity as measured in rat H-4-II E hepatoma cells (Hanberg et al., 1991).Using a recombinant H-4-II E rat hepatoma cell line having Ah-receptormediated expression of a luciferase reporter gene (the CALUX as-say)(Aarts et al., 1995; Murk et al., 1998), a number of individual PBDEcongeners have been tested for their potency to activate/deactivate the Ahreceptor (Meerts et al., 1998a). In order to study antagonism, the samePBDE congeners were also tested in the presence of 2,3,7,8-tetrachloro-dibenzo-p-dioxin (TCDD). Seven of the 17 PBDE congeners (BDEs -32,-85, -99, -119, -153, -166, -190) tested showed ability to activate the Ah-receptor. Potencies could only be determined for BDE-166 and BDE-190and these are in the same range as the mono-ortho PCB congeners 105and 118 (Sanderson et al., 1996). Some congeners such as BDE-85, -99and -119 showed both agonist and antagonist activities depending on theconcentration tested. Nine congeners, including BDEs -15, -28, -47, -77and -138, showed antagonist activities against TCDD. The observed an-tagonism may be due to competition between PBDEs and TCDD at theAh-receptor level.

Studies of the potency of 17 PBDE congeners and their hydroxylatedmetabolites for competitive binding to human transthyretin in vitro haverecently been carried out (Meerts et al., 1998b). Transthyretin is the thyroidhormone transport protein present in plasma and has a binding site for thethyroid hormone thyroxine (T4). The parent compounds were incubatedwith rat hepatic microsomes from rats treated with beta-naphthaflavone(NF), phenobarbital (PB) or clofibrate (CLOF). The parent compounds andthe metabolites formed from the different microsomal incubations werethen tested for their ability to compete with T4 for binding to transthyretin.The results are given in Table 4 and show that a number of metaboliteswere potent competitors for transthyretin. The parent compounds showed

32

no competition. For example, after PB-microsomal incubation of 2,2',4,6'-tetraBDE (BDE-51), hydroxylated PBDE metabolites displaced T4 fromtransthyretin with fairly high potency (Meerts et al., 1998b). The results in-dicate that hydroxylated metabolites of PBDE may be potent competitorsof T4 and could disrupt normal thyroid hormone function in wildlife andhumans if present.

Table 4. Inhibition of T4-transthyretin binding in vitro by PBDE metabolites obtained afterincubation with rat hepatic microsomes induced by phenobarbital (PB), beta-naphthafla-vone (NF) or clofibrate (CLOF). Inhibition potencies are given from the undiluted extract.++=60% inhibition, +=20% inhibition, – =0–20% inhibition. (From Meerts et al., 1998b).

Transthyretin carries T4 in the plasma to the target tissues, where T4 isthen deiodinated to 3,3',5-triiodothyronine (T3) (Figure 2). T3 then inter-acts with two sub-types of thyroid hormone receptors (THRs) designatedalpha and beta. The T3-THR complex can then bind to response elementson the DNA that regulate the transcription of thyroid hormone activated

BDE congenertested

PB microsomes NF microsomes CLOF microsomes

15 ++ ++ –

28 ++ + +

30 ++ ++ ++

32 – + +

47 ++ – –

51 ++ – +

71 + + +

75 ++ + +

77 ++ + +

85 + – –

99 + – –

100 ++ – –

119 ++ – –

138 – – –

153 – – –

166 + – –

190 – – –

33

genes. T4 can also interact with THR but has only about 10% of the poten-cy of T3. To determine the potency of hydroxylated PBDE to bind to THR-alpha and THR-beta, several brominated structural analogues of T4 and T3were synthesized: 4-hydroxydiphenyl ether, 4-hydroxy-2',4',6'-tribro-modiphenyl ether (III), 3-bromo-4-hydroxy-2',4',6'-tribromodiphenyl ether(IV) and 3,5-dibromo-4-hydroxy-2',4',6'-tribromodiphenyl ether (V) (Fig-ure 5) (Marsh et al., 1998).

The results showed that the highest affinity was found for the T3 ana-logue (IV), with the T4 analogue (V) showing about one third of the affin-ity of the T3 analogue. The affinity of analogue III was low and was lowestfor 4-hydroxydiphenyl ether (Marsh et al., 1998). The brominated ana-logues had lower affinities than T3 and T4, probably because of the lack ofthe 4-carboxyl group. The results indicate that hydroxylated metabolites ofPBDE may not only disrupt the normal transport of T4 to target tissues, butmay also be able to bind to the thyroid hormone receptors, thus influencingthe regulation of thyroid hormone dependent genes within the cell nucleus.

Mitogen-induced DNA synthesis and immunoglobulin synthesis by hu-man lymphocytes in vitro was examined after exposure to purified BDE-47 and -85. No effects on mitogen-induced proliferation or immunoglobu-lin synthesis were observed (Fernlöf et al., 1997). The results indicate thatproliferation and immunoglobulin synthesis are insensitive to the direct ac-tion of polybrominated diphenyl ethers. Exposure to different PCB conge-ners (CB-77, CB-118, CB-153) also gave no effects.

BDE-47 was shown to induce a statistically significant increase in intra-genic recombination when studied in one of two tested in vitro assays usingmammalian cells (Helleday et al. 1999). This may indicate that BDE-47

Figure 5. Synthetic pathway and chemical structures of several hydroxylated PBDE (Marsh et al., 1998).

34

can induce cancer via a non-mutagenic mechanism, similarly to other en-vironmental contaminants such as DDT and PCB.

4.2.1.2 In mammalsTechnical PBDE products are able to induce both phase I and phase II de-toxification enzymes in the liver. Regarding the cytochrome P450 (CYP)mediated phase I metabolism, CYP1A1 and 1A2 are induced; this could beshown by the increased activity of liver microsomal 7-ethoxyresorufin O-deethylase (EROD) after Bromkal 70 (a commercial pentaBDE) exposurein Wistar rats (von Meyerinck et al. 1990) and in H-4-II E cells (Hanberget al., 1991). Other enzymes that are used as indicators of microsomalphase I activity were also induced by PBDEs (technical pentaBDE prepa-rations in rats) including benzphetamine N-demethylation, p-nitroanisoledemethylase, aryl hydrocarbon hydroxylase (AHH) and benzo(a)pyrenehydroxylase (von Meyerinck et al., 1990; Carlson 1980a; b). Some of theenzymes were induced in a long-term oral administration study in rats at aconcentration as low as ca 1 µmol/kg (technical pentaBDE), and the en-zymes remained induced 30–60 d after termination of exposure (Carlson,1980b). DeBDE seems to have low enzyme-inducing potency. However,as CYP1A1 and 1A2 are typically induced by halogenated dioxin-likecompounds, possible contaminants with Ah-receptor binding affinitypresent in technical PBDE mixtures could be responsible for the enzymeinduction seen in these studies.

In interaction studies of PBDE, PCB and chlorinated paraffins (CP), mi-crosomal enzyme activities were studied in rats (Hallgren and Darnerud,1998). The results showed that the pure congener BDE-47 (2,2',4,4'-tetra-BDE) increased EROD and MROD activities only somewhat (to twice thecontrol levels), indicating that this substance has a low CYP1A1/2 induc-ing activity. In the same study, PCB (Aroclor 1254) markedly inducedEROD and MROD. Earlier results on a commercial PBDE mixture (Bro-mkal 70) showed a rather strong induction of EROD and MROD in rats,which suggests that the Bromkal mixture contained CYP 1A1/2-inducingsubstances, probably present as contaminants (unpublished studies inDarnerud and Sinjari, 1996).

Regarding other microsomal enzymes, PROD levels (which mirrorCYP2B activities) were measured in rats exposed to BDE-47 (Hallgren andDarnerud, 1998). In this case the PROD levels were dose-dependently el-

35

evated up to 10 times in exposed animals, and the levels were about thesame as those found after PCB exposure. As the microsomal enzyme in-duction results are indicative of operational metabolic systems, PBDE maytherefore, at least to some extent, be transformed by CYP2B in the phase Imetabolism step.

In studies on phase II induction, three different PBDE fractions, i.e. alow (24% tetra, 50% penta) and a high (45% hepta, 30% octa) brominatedmixture, and the DeBDE congener only, were tested. Daily oral adminis-trations (14 days, 0.1 mmol/kg body wt.) of both mixtures, but not DeBDE,resulted in a long-lasting induction of uridine diphosphate glucuronyltransferase (UDPGT) activity in rats (Carlson, 1980a).

Short-term feeding studies using PeBDE in rats led to changes in the liv-er (increased weight, hepatocytomegaly) and thyroid (hyperplasia) (Norriset al., 1975b; El Dareer et al., 1987; Great Lakes Chem. Corp, undated a).No carcinogenicity studies have been performed for TeBDE or PeBDE.

The Bromkal 70-5DE product, containing mainly Te- and PeBDE, caus-es decreased thymus weight, increased liver/body weight ratios in mice anddecreases in the thyroid hormone thyroxine in rats and mice (Fowles et al.,1994; Darnerud and Sinjari, 1996). Decreases in thyroxine were also seenwhen rats and mice were treated with the single congener BDE-47, but noeffects on thyroid stimulating hormone were seen for BDE-47 or Bromkal70 (Darnerud and Sinjari, 1996).

In subsequent studies the interactive effects of different organohalogencompounds (PCB, PBDE and CP) on thyroxine hormone levels and micro-somal enzyme activities were tested (Hallgren and Darnerud, 1998). Fe-male rats were orally exposed to single compounds or combinations dailyduring 14 days. The results show that PCBs (Aroclor 1254) and PBDEs(BDE-47) significantly reduce the T4 levels in rats, in the actual exposureinterval (6–18 mg/kg body weight/day), and that Aroclor 1254 results inthe strongest effect, when administering the substances orally in isomolarconcentrations. EROD and MROD, but to a lesser extent PROD and UD-PGT, activites correlated to T4 effects, which could indicate that glucuro-nidation of T4 is not a major factor in explaining the observed decrease inT4 plasma levels. Regarding the mixed BDE-47 + CP group, a synergisticdecrease in free T4 levels, and increase in EROD activity, was observed.As organisms are exposed to these environmental chemicals as mixtures,the observed interactive effects are of interest.

36

It is known that hydroxylated metabolites of PCB can compete with thy-roxine for the binding site on the thyroxine-carrying protein transthyretinin plasma (Brouwer et al., 1990). The effects on the thyroid gland (hyper-plasia, decreased thyroxine levels) seen with PBDE could also be due to ef-fects of hydroxylated metabolites, which have been found as metabolitesin both mice, rats and fish. Some hydroxylated PBDEs have structural sim-ilarities to thyroxine. In preliminary experiments, metabolites formed in ratliver from BDE-47, -99 and -153 have been shown to be better competitorsfor the transthyretin binding site than the parent compounds, which indi-cates they could compete with thyroxine as well (Brouwer and Murk, per-sonal communication, cited in Örn, 1997).

Short-term feeding studies with high doses of DeBDE in rats led to liverlesions and thyroid hyperplasia. Long-term exposure to DeBDE was alsofound to induce thyroid hyperplasia, hepatocellular and thyroid adenomasand carcinomas in mice (Great Lakes Chem. Corp, undated b; 1987).Workers producing DeBDE and decabromobiphenyl had a statistically sig-nificant increase in hypothyroidism (Bahn et al., 1980).

Immunotoxicity was studied after oral treatment with Bromkal 70-5 DEor BDE-47 in rats and mice (Darnerud and Thuvander, 1998). In mice,BDE-47 caused reduced splenocyte number as reflected in decreased num-bers of CD45R+, CD4+ and CD8+ cells in spleens. In mice treated withBromkal 70-5 DE, absolute numbers of double negative thymocytes weresignificantly lower than in controls and mice also showed reduced produc-tion of IgG. No effects were seen in rats. Thus, BDE-47 and Bromkal 70-5 DE, which contains BDE-47, both seem to be immunotoxic in mice.

Studies have shown that there is a critical phase in neonatal mouse braindevelopment when the brain is particularly susceptible to effects of low-dose exposure to toxic substances such as PCB, DDT, pyrethroids, organo-phosphates, paraquat and nicotine (Eriksson, 1997). This critical phase isknown as the ”brain growth spurt” (BGS) and disruption leads to persistentdisruption in adult brain function. The BGS occurs at different time pointsin different mammalian species (Davison and Dobbing, 1968). In rats andmice it occurs in the first 3-4 weeks of life (neonatal period) whereas in hu-mans it occurs during the third trimester of pregnancy and throughout thefirst 2 years.

To study possible neurotoxicity of brominated flame retardants, BDE-47, BDE-99 or TBBPA were administered orally to neonatal mice on day

37

10 (Eriksson et al., 1998). Doses administered are given in Table 5. Severaltests of behavior, locomotion, activity and memory were performed withthe treated individuals several months later. Results showed that BDE-47and -99 both induced permanent aberrations in spontaneous motor behav-ior which worsened with age (Table 5). Similar effects have been seen inmice exposed neonatally to similar molar doses of some ortho-substitutedPCB and co-planar PCB (Eriksson et al., 1991; Eriksson and Fredriksson,1996a; 1996b; 1998). Neonatal exposure to BDE-99 also affected learningand memory functions in the adult animal. No effects were seen forTBBPA.

Table 5. Neurotoxicological effects seen in mice several months after administration of asingle dose of specific PBDE congeners and TBBPA on day 10 after birth. Results fromEriksson et al., 1998).

+ indicates permanent aberrations; – indicates no effects

In a follow-up study, Eriksson et al. (1999) investigated whether there isa critical time in neonatal mouse brain development for induction of theneurotoxic effects of BDE-99. One single oral dose of 8 mg/kg bodyweight (14 µmol/kg body weight) was administered to 3-day, 10-day and19-day-old mice. Spontaneous behavioral tests were performed after 4months. The mice exposed to BDE-99 on day 10 showed significant behav-ior aberration, as was previously seen, and mice exposed on day 3 showedsimilar aberration but to a lesser degree. The mice exposed on day 19showed no significant change from the controls.

Uptake and retention of BDE-99 in the brain was also studied by admin-istering 14C-labelled BDE-99 to 3-day, 10-day and 19-day-old mice (Eriks-son et al., 1999). The amounts of radioactivity found in the brain were

Substance Dose in mg/kg body weight

Dose in µmol/kg body weight

Spontaneous motor behavior disfunction

Learning and memory disfunction

BDE-47 low 0.7 1.4 – –

BDE-47 high 10.5 21.1 + –

BDE-99 low 0.8 1.4 + –

BDE-99 high 12.0 21.1 + +

TBBPA low 0.75 1.4 – –

TBBPA high 11.5 21.1 – –

38

measured at 24 h and 7 days after administration. The retention of BDE-99was similar to what has been observed after neonatal exposure to CB 52,CB 153 and DDT (Eriksson, 1997). The retention of BDE-99 in mice ex-posed on day 3 indicates that the effects seen may be due to the amount stillpresent in the brain on day 10. The neurotoxic effects seem to involvechanges in the cholinergic system as mice given BDE-99 on day 10 andthen challenged as adults with a low dose of nicotine behave completelythe opposite of controls. From these studies, it was concluded that the win-dow for permanent effects of BDE-99 and BDE-47 is day 10 in neonatalmice (Eriksson et al., 1998; 1999).

Female rats given BDE-47 orally for a period of two weeks were thenkilled and the choroid plexus of the brain removed, homogenized and in-cubated with 125I-T4 (Sinjari et al., 1998). Compared to controls, there wasa dose-dependent reduction in the binding of 125I-T4 to the choroid plexus.In contrast, in vitro incubation of rat choroid plexus with BDE-47 revealedno competitive inhibition of labelled T4 binding. This indicates that BDE-47 metabolites can cross the blood-brain barrier and bind to the choroidplexus T4-binding sites. This in turn could cause the interference of T4transport to the brain, with risks for effects on neural development.

4.2.1.3 In fish Microinjection of Bromkal 70-5DE in rainbow trout larvae led to weaklyinduced EROD activity (Norrgren et al., 1993). Three-spined sticklebackfed Bromkal 70-5DE showed weakly induced liver EROD activity, fattylivers and a reduction in spawning success (Holm et al., 1993). Microinjec-tion of 2,2',4,4'-TeBDE, 2,2',3,4,4'-PeBDE or 2,2',4,4',5-PeBDE into new-ly fertilized rainbow trout eggs in an early life stage mortality bioassayshowed no effects compared to 2,3,7,8-tetrachlorinated dibenzo-p-dioxin(Hornung et al., 1996). However, 3,3',4,4'-tetrabrominated biphenyl and3,3',4,4',5,5'-hexabrominated biphenyl were ten-fold more potent than theidentically chlorinated biphenyls.

In a recent study, rainbow trout were fed food containing either 2,2',4,4'-TeBDE (BDE-47) or 2,2',4,4',5-PeBDE (BDE-99) for 6 and 22 days tostudy biological effects. Both congeners were found to significantly inhibitEROD activity in the liver with BDE-47 being most powerful in this effect(Tjärnlund et al., 1998). A 35% reduction of GSH-reductase was also ob-served after 22 days. Hematocrit and blood glucose also showed small but

39

statististically significant changes after 6 days. No differences were seen incondition factor, liver somatic index, spleen somatic index, numbers ofleucocytes, thrombocytes, granulocytes, lymphocytes or hemoglobin lev-els. When injected into fertilized fish eggs, no effects were seen on thia-mine use.

4.2.2 TBBPA

4.2.2.1 IN VITRO

Studies of the potency of TBBPA for its competitive binding to humantransthyretin in vitro have shown that TBBPA has the highest potency ofall brominated and chlorinated substances tested so far. It is up to 25 timesmore potent than T4 (Brouwer, 1998). However, this effect was not seen inin vivo studies in pregnant rats, see below (Meerts et al., 1999).

TBBPA was tested for its activity in inducing intragenic recombinationin two in vitro assays using mammalian cells but caused no effect (Helle-day et al., 1999).

4.2.2.2 In mammalsTBBPA fed orally to mice and rats showed low or no effects on behavior,weight gain, mortality, organ abnormalities or hematology (WHO/IPCS,1995).

The thyroid transport protein, transthyretin, may play a major role in de-livering T4 from the mother to the fetus across the placental barrier as wellas across the blood-brain barrier, where it is converted to T3, an essentialhormone for normal brain development (Calvo et al., 1990; Southwell etal., 1993). TBBPA has been shown to competitively bind to human tran-sthyretin in vitro with high affinity. To study the effects of TBBPA on thy-roid hormone transport, including across the blood-brain barrier, pregnantrats were given 14C-labelled TBBPA on days 10 to 16 of gestation (Meertset al., 1999). Blood plasma thyroid hormone and thyroid stimulating hor-mone levels were measured as well as 125I-T4-competition binding on ma-ternal and fetal plasma transthyretin. No effects were seen on total T4, freeT4 or total T3 levels in dams or fetuses. Thyroid stimulating hormone in-creased significantly in fetal plasma by 196% but no effect was seen indams. No 14C label was detected on transthyretin and there was no shift inbinding of 125I-T4, which would be expected if TBBPA had bound to trans-

40

thyretin. Therefore, TBBPA was concluded to not bind to transthyretin invivo. No selective accumulation of TBBPA-related activity was found inthe fetal brain.

4.2.2.3 In fishBluegill sunfish exposed to TBBPA in water became irritated and exhibit-ed abnormal swimming behavior. Rainbow trout exhibited irritation,twitching, erratic swimming, dark discoloration and labored respiration.Fathead minnow showed reduced survival of young at hatch and reducedsurvival and growth after 30 days (WHO/IPCS, 1995). There is no infor-mation regarding uptake, metabolism or effects in vitro, in mammals or infish of the methylated derivative of TBBPA (MeTA).

4.2.3 HBCD

4.2.3.1 IN VITRO

The effects of HBCD on intragenic recombination were studied in two invitro assays using mammalian cells (Helleday et al. 1999). HBCD causedstatistically significant increases in recombination frequency in both testsystems, indicating that it may induce cancer via a non-mutagenic mecha-nism, similarly to other environmental contaminants such as DDT andPCB.

4.2.3.2 In mammalsHBCD is not acutely toxic to rats or mice when given orally or in rats wheninhaled. No acute dermal toxicity is seen in treated rabbits. Chronic expo-sure in rats leads to increased liver weights, and in one study, thyroid hy-perplasia. Chronic exposure in female rats led to inhibited oogenesis (Zel-ler and Kirsch, 1969). In chronic oral studies in pregnant rats, HBCD wasfound to suppress maternal food consumption and increase maternal liverweight in the highest dose group (1% of diet), but no effects were seen onthe offspring (IUCLID, 1996). The quality of most of these studies is, how-ever, under question.

4.2.3.3 In fishNo information is available on the toxicity of HBCD to fish.

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5. Environmental concentrationsA summary of environmental concentrations is given in Table 17 at the endof this section.

5.1 Abiotic samplesProductsIn laboratory experiments to test if all TBBPA is actually bound, filingsfrom a circuit board were extracted with solvent and analyzed and the re-sults showed that a small percentage of the TBBPA was not bound (Sell-ström and Jansson, 1995). Thus TBBPA can also leak from products intothe environment.

Qualitative analyses have shown the presence of HBCD in styrofoamchips used as packing material (Amelie Kierkegaard and Ulla Sellström,unpublished results).

5.1.1 AIR

In 1979, DeBDE was identified in air particulate in the vicinity of plantsmanufacturing brominated flame retardants (Zweidinger et al., 1979). Wa-tanabe and his coworkers found predominantly DeBDE in airborne dustfrom the Osaka region, in Japan (Watanabe et al., 1995). In samples fromTaiwan and Japan in the vicinity of metal recycling plants, various tri-, tet-ra-, penta- and hexaBDEs were detected in air (Watanabe et al., 1992).Concentrations ranged from 23–53 pg/m3 in Taiwan and 7.1–21 pg/m3 inJapan.

One explanation for the wide distribution of lower brominated PBDEsfound in the Swedish environment could be long range transport in air. Totest this, two high-volume (48-hour) air samples previously collected with-in the Swedish Dioxin Survey, one from Hoburgen in southern Sweden(Gotland) and one from Ammarnäs in northern Sweden, were analyzed.The sampling conditions for these two samples are given in Table 6.

The results showed quantifiable amounts of BDE-47, -99 and -100 andHBCD in both samples (Bergander et al., 1995). The total PBDE levels

42

were approximately 1 and 8 pg/m3. Highest levels of BDE-47 were foundon polyurethane foam plugs (gas phase) while higher levels of BDE-99 and-100 and HBCD were found on filters (particulate phase) (Figure 6). Thiscorresponds well with the physicochemical properties of these substances,as BDE-47 is more volatile than the other compounds, and would be ex-pected to be in the gas phase to a larger extent. No DeBDE was found, butthe detection level for DeBDE is much higher than for the lower brominat-ed PBDEs. Because the samples were taken at different times of year andthe trajectories of the air packets were not stable over the sampling period,it is not possible to compare the concentrations between the two sites.However, the fact that these substances were found in air samples fromboth sites supports the hypothesis that these substances are spread via long-range transport.

Support for this theory has also come from newer results. Air sampleswere collected at one rural site in southern England (Stoke Ferry) and one

Table 6. Sampling conditions for two high-volume air samples.

Site Volume sampled Temperature °C Sampling date

Ammarnäs 1475 m3 –4 January, 1991

Hoburgen 1736 m3 +15 July, 1990

0

1

2

3

4

5

6

PUF Filter PUF Filter

Ammarnäs Hoburgen

pg

/m3

BDE-47BDE-99BDE-100HBCD

Figure 6. Approximate concentrations of PBDE and HBCD in air samples fromtwo sites in Sweden (Bergander et al., 1995).

43

semi-rural site in northwestern England (Hazelrigg) during 1997 (spring,summer, autumn and winter) and analyzed for PBDEs (Peters et al., 1999a;b). Detectable concentrations of tri- to heptaBDEs were found and the sumconcentrations of BDE-47, -99 and -100 were 7–28 pg/m3 at Hazelrigg and11–67 pg/m3 at Stoke Ferry (A. Kierkegaard, Stockholm University, per-sonal communication). PBDE has also been measured in several archivedair samples from the Arctic (Alert, Northwest Territories, Canada andDunai Island, eastern Siberia) taken between January 1994 and January1995 (Alaee et al., 1999 and M. Alaee, Dept. of Environment, NWRI, Can-ada, personal communication). The sum concentrations of several di- tohexaBDEs were 1–4 pg/m3 at Alert most of the year, but 28 pg/m3 in July,1994. Sum concentrations in air samples from Dunai, were somewhat low-er than at Alert, with the highest level also found in summer (7–8 pg/m3).BDE-47 and -99 were the major congeners found.

Bergman et al. (1997) developed a sampling technique for sampling airparticulates in the working environment. They determined a number ofbrominated flame retardants on air particulates in rooms containing com-puters and other electrical equipment. All particulate samples containedTBBPA, BDE-47 and BDE-99. The presence of these substances in air par-ticulates shows that these substances are leaking into the indoor environ-ment from electronic devices and therefore exposing humans.

In another study, air was sampled in an electronics dismantling plant, inan office with computers and outdoors (Bergman et al., 1999; Sjödin et al.,1999b) and analyzed for several PBDE as well as TBBPA. Highest concen-trations of all substances were found in air from the electronics dismantlingplant. Concentrations in the office were not detectable or 400-4000 timeslower and not detectable concentrations were found for the outdoor air.Mean concentrations found in air at the dismantling plant were 2.5 pmol/m3 (1250 pg/m3) for BDE-47, 4.6 pmol/m3 (2600 pg/m3) for BDE-99, 6.1pmol/m3 for BDE-153, 26 pmol/m3 for BDE-183, 38 pmol/m3 (36500 pg/m3) for BDE-209 and 55 pmol/m3 (29900 pg/m3) for TBBPA (Bergman etal., 1999). Samples were also taken near a plastics shredder at the disman-tling plant to identify possible point sources (Sjödin et al., 1999b). Concen-trations of the PBDEs were found to be 4–10 times higher in proximity ofthe shredder when compared to the air samples at other sites in the dismant-ling plant.

44

5.1.2 WATER

Plastic parts and textiles are flame-retarded in cars. Scrapped cars are takenapart at car dismantling centers, so that parts that can be recycled are re-moved. However, some materials, possibly containing PBDEs, cannot bereused and are placed in landfills. A car parts removal center in the city ofHalmstad has a landfill where leach water is collected in a lagoon for sed-imentation. The water eventually is led to the sewage treatment plant inHalmstad (see sewage sludge results below).

In order to see if PBDEs, especially DeBDE, are leaching from the land-fill, samples of incoming leach water to the lagoon as well as lagoon sludgewere collected and analyzed for PBDEs. This was the first attempt to ana-lyze a water sample for these substances, which required some methods de-velopment at the extraction stage. The results are preliminary and lowestpossible concentrations (Amelie Kierkegaard, Stockholm University, per-sonal communication; de Wit, 1995; 1997).

The results showed low PBDE concentrations in both the water andsludge samples. The leach water contained 0.3 ng/liter (sum of BDEs -47,-99 and -100) and the sludge, 2 ng/g dry weight (sum of BDEs -47, -99 and-100). Neither sample contained octa-, nona- or DeBDE. Therefore, thisparticular industry does not seem to be the source of the high DeBDE con-centrations in the sewage sludge from Halmstad.

5.1.3 SEWAGE SLUDGE

Sewage sludge samples collected in 1988 from Ryaverket sewage treat-ment plant in Gothenburg, Sweden, were analyzed for PBDEs. Concentra-tions of BDE-47, -99 and -100 together were around 20–30 ng/g dry weight(Nylund et al., 1992) with BDE-47 and -99 present at similar levels. Theproportions seen in sewage sludge are similar to those seen in the technicalPeBDE product Bromkal 70-5DE. The levels were similar to those foundin Germany by Hagenmeier et al. (1992), who found tri- to heptaBDEs insewage sludge, with concentrations of Te- and PeBDEs (not designatedwhich congeners) of 0.4–15 ng/g.

Several digested sewage sludge samples were collected in Sweden in1991 within a project administrated by the Swedish EPA. The sampleswere taken from the sewage treatment plant at Halmstad. The purpose ofthe project was to analyze sewage sludge for a number of organic environ-

45

mental toxins including PBDEs and to study what happens to these sub-stances when the sludge is plowed into soil as fertilizer.

One of these sludge samples was analyzed for PBDEs and resultsshowed that the sum of concentrations of BDEs -47, -99 and -100 was ap-proximately at the ppb level (ng/g), similar to the levels found in a sampleanalyzed previously from Ryaverket (Nylund et al., 1992). However, theconcentration of DeBDE was 1000 times higher than for other PBDEs, atthe ppm level (µg/g). This sample also contained high concentrations ofnona- and octaBDEs (A. Kierkegaard, personal communication; de Wit,1995).

In 1988, a sewage sludge sample was collected from a treatment plantreceiving leach water from a landfill where wastes from a plastics industryare placed (Klippan). This plastics industry uses TBBPA. For comparison,a second sewage sludge sample was collected from a treatment plant (Rim-bo) having no known sources of TBBPA connected to it (Sellström andJansson, 1995). The samples were analyzed for TBBPA and MeTA and re-sults showed that MeTA was not detectable, but TBBPA levels were 56and 31 ng/g dry weight. The samples were also analyzed for BDE-47, -99and -100. The sum of these three PBDE congeners in the Klippan samplewas 45 ng/g dry weight and for the Rimbo sample, 119 ng/g dry weight(Sellström, 1999).

Sewage sludge samples from sewage treatment plants in Stockholm,Sweden, have recently been analyzed for TeBDE, PeBDE, DeBDE, HBCDand TBBPA (Sellström et al., 1999; Sellström, 1999). The results are givenin Table 7 below. The concentrations of BDE-47, -99, -100 and -209 do notdiffer as much between plants as do TBBPA and HBCD.

Table 7. Mean concentrations of several PBDE (n=4), HBCD (n=4) and TBBPA (n=2) insewage sludge from three treatment plants in Stockholm in ng/g dry weight.

5.1.4 SEDIMENT Previous studies in Japan have found TeBDE, PeBDE, HxBDE andDeBDE in river sediments (Watanabe et al., 1986; 1987; 1995). Concen-

Treatment plant BDE-47 BDE-99 BDE-100 BDE-209 HBCD TBBPA

Sickla 78 98 24 220 21 3.6

Bromma 80 100 25 270 54 8.6

Loudden 36 56 13 170 19 45

46

trations of TeBDE and PeBDE together were 21–59 ng/g dry weight.DeBDE was found in concentrations ranging from <25 ng/g to 11600 ng/gdry weight (Environmental Agency Japan, 1983; 1989; 1991).

The upper layer of a sediment core collected in the southern Baltic Sea(Bornholm Deep) was analyzed for BDEs -47, -99 and -100 and contained0.52 ng/g dry weight (2.9 ng/g ignition loss) (sum of three congeners)(Ny-lund et al., 1992). In a more recent study, twenty surficial sediment sam-ples taken from numerous sites in the Baltic Sea during 1993 were ana-lyzed for BDEs -47, -99 and -100 within a Helsinki Commission (HEL-COM) sediment baseline study (Jonsson and Kankaanpää, 1999). Resultsshowed low levels in all samples, from not detected to 5.4 ng/g ignitionloss based on the sum of three congeners (Table 8).

Table 8. Concentrations of 2,2'4,4'-TeBDE (BDE-47), 2,2',4,4',5-PeBDE (BDE-99) and2,2',4,4',6-PeBDE (BDE-100) in sediment samples from several sites in the Baltic Sea.Concentrations are given in ng/g ignition loss (ng/g IG) and all samples were analyzed induplicate. < means not detected at this level.

Sample site Ignition loss % BDE-47 BDE-99 BDE-100

Bothnian Sea 8 <1.6 1.2 *

Gulf of Finland 10 1.4–1.9 1.1–1.9 0.36–0.48

Gulf of Finland 13 <1.4 <1.2 <0.33

Gulf of Finland 6 <1.9 <1.6 <0.43

Gulf of Finland 10 <1.7 <1.4 <0.38

Gulf of Finland 16 1.4–1.9 1.1–1.2 0.32

Bothnian Bay 12 <1.6 <1.1 *

Åland Sea 9 1.6–1.7 1.2–2.4 1.3

Northern Baltic Proper 21 <1.1 <1.0 <0.28

Bothnian Bay 9 <1.8 <1.2 *

Kieler Bight 13 3.2–3.4 0.63–0.72 0.48–0.70

Lithuanian Coast 13 1.8–2.3 1.7–2.2 0.37–0.53

Gotland Deep 8 <1.4 <1.3 <0.33

Gdansk Bay 15 1.6 1.3 0.31

Härnösand 8 <2.0 <1.4 *

Kattegatt 5 <1.4 0.55-0.76 *

Bornholm Deep 14 <1.1 0.79-85 0.24

Riga Bight 3 15 0.85-1.0 0.69-0.76 0.19-0.21

47

* Not quantified due to interfering substances

In a study near a Swedish plastics industry using TBBPA, sediment sam-ples were collected up- and downstream of the industry and these were an-alyzed for TBBPA and MeTA (Sellström and Jansson, 1995) as well as forBDE-47, -99 and -100 (Sellström, 1996; 1999). TBBPA, MeTA and allthree PBDE congeners were found in higher concentrations downstream ofthe plant than upstream (Table 9). This indicates that the plastics industrywas the most likely source for these substances.

Table 9. Concentrations (ng/g ignition loss) of TBBPA, MeTA and BDEs -47, -99 and -100in sediments taken upstream and downstream of a plastics industry (Sellström and Jansson,1995).

Surficial sediment samples were collected in 1995 at 8 sites along theRiver Viskan, where numerous textile industries are located (Figure 7).These industries have used various brominated flame retardants in produc-tion of textiles. BDEs -47, -99, -100 and -209 as well as HBCD were quan-tified in the sediments and the concentrations increased further down-stream as more industries were passed (Sellström et al., 1998b). The con-centrations of BDEs -47, -99 and -100 together ranged from not detectedto120 ng/g ignition loss, BDE-209 ranged from not detected to 16 000 ng/g ignition loss and HBCD ranged from not detected to 7600 ng/g ignitionloss (Figure 8). The lowest levels of the PBDEs and HBCD were found up-stream of the industries. This is the first time HBCD has been found in en-vironmental samples in Sweden.

A study of various contaminants in sediments collected from the mouthsof major European rivers included several brominated flame retardants(Kierkegaard et al., 1996; van Zeijl, 1997; Sellström et al., 1999). Resultsare shown in Figure 9. High levels of BDE-47 and -99 were found in two

Riga Bight 5 15 <0.99 0.67 <0.18

East Gotland 35 2.1 1.5 0.32

Site Ignition loss %

TBBPA MeTA BDE-47 BDE-99 BDE-100 Sum of 3 PBDEs

Upstream 67 50 36 3.7 8.8 1.6 14.1

Downstream 62.5 430 2400 780 1200 270 2250

Sample site Ignition loss % BDE-47 BDE-99 BDE-100

48

rivers in Great Britain (Humber and Mersey) and two in the Netherlands(sum of two congeners 1.61–13.1 ng/g dry weight). Highest 2,2',4,4',5,5'-HxBB levels were found in the Seine (France), three rivers in the Nether-lands, the rivers Schelde (Belgium), Forth (Great Britain) and Ems (Ger-many) (range 0.013–0.056 ng/g dry weight). Levels of DeBB were highestin sediment from the Seine (2.4–3.9 ng/g dry weight). DeBDE (BDE-209)levels were highest in the River Mersey (Great Britain), followed by theSchelde and River Liffey (Ireland) (range 34–1800 ng/g dry weight).

Allchin et al. (1999) have recently carried out a survey of PBDE in sed-iments and fish (see section 5.2.2.1) from several rivers and estuaries in theUnited Kingdon. Sediments were collected upstream and downstream ofsuspected sources including a manufacturer of PeBDE and OcBDE, sever-al industries using PeBDE, several landfills receiving wastes suspected tocontain PBDEs and a reference site (Table 10).

Figure 7. Map showing sampling sites for sediments along the River Viskan.

49

The highest concentrations of BDE-47, -99, PeBDE (DE-71) andOcBDE (DE-79) were found in sediments around or downstream of themanufacturing site at Newton Aycliffe on the River Skerne (Table 11). Thehighest concentrations of DeBDE (DE-83) were found downstream of asewage treatment plant on the River Calder, but high concentrations werealso seen on the River Skerne downstream of the manufacturing site. Thegeneral conclusions drawn from this study were that the PeBDE andOcBDE manufacturing plant at Newton Aycliffe is a major source of PBDEon the River Tees/Skerne including the mouth of the river, which is 40 kmdownstream of the plant. Other sources are also implicated along other UKrivers although these could not always be identified. BDE-99 concentra-tions were similar or somewhat higher than for BDE-47 in most sediments.

02000400060008000

1000012000

1 2 3 4 5 60

10

20

30

40

50

1 2 3 4 5 6

BDE47 in Sediments

Sampling Location Sampling Location

ng/g ig.

BDE209 and HBCD in Sedimentsng/g

ig.

Figure 8. BDE-47, BDE-209 and HBCD concentrations in sediments fromRiver Viskan (Sellström et al., 1998b).

50

0

1

2

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7

8

Sei

ne

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de

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ber

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sey

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Tha

mes

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NZ

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S10

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W

WW

Z

NZ

RN

W2

NZ

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S

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elde

Elb

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Ski

en

Glo

mm

a

Got

a

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ng

/g (

dry

wt)

HxBBTeBDEPeBDEDeBBDeBDE/100

17

Figure 9. Mean concentrations (ng/g dry weight) of HxBB, DeBB, BDE-47, sum of BDE-99 and -100, and BDE-209 in sedi-ments from the mouths of several European rivers (Kierkegaard et al., 1996; Sellström et al., 1999). SHW = Southampton,NZRTS100 = North Sea (reference site), NWW = Rhine, WWZ = Wadden Sea, NZRNW2 = Nordwijk, Netherlands and NZBCS= North Sea, Belgium.

51

Table 10. Sampling sites in the United Kingdom with description of possible sources (fromAllchin et al., 1999).

Table 11. Concentrations of BDE-47, -99, PeBDE quantified as the technical product DE-71,OcBDE quantified as the technical product DE-79 and DeBDE quantified as the technicalproduct DE-83 in sediments from UK rivers (ng/g dry weight). Data from Allchin et al. (1999).

Site Description

River Tweed Reference site

Rivers Skerne and Tees

Manufacturer of PeBDE and OcBDE at Newton Aycliffe (on River Skerne). The River Skerne is a tributary of the River Tees.

Rivers Calder and Ribble

Foam manufacturer using PeBDE at Accrington.

River Nith Rubber and tires manufacturer using PeBDE.

Avonmouth and Bristol Channel

Main UK user of PeBDE.

Great Ouse Elstow landfill site which receives waste from the manufacturer at Newton Aycliffe.

Leeds and River Humber

Landfill site receiving car waste, possibly generating leachate containing PeBDE.

Site BDE-47 BDE-99 DE-71 (PeBDE)

DE-79 (OcBDE)

DE-83 (DeBDE)

River Tweed <0.3-0.4 <0.6 <0.38 <0.44 <0.6

River Tees, upstream <0.3 <0.6 <0.38 <0.44 <0.6

River Tees/Skerne, downstream of confluence

8–51 11–85 34–35 25–129 <0.6–7

River Skerne, Newton Ay-cliffe

239 319 130 397 64

River Skerne, downstream Newton Aycliffe

68–112 111–159 45–68 264–1405 23–294

Tees estuary 8.9–368 16–898 19–366 29–1348 <0.6–9

River Calder, upstream of sewage plant

2.3–7.6 0.6–16 <0.38–6.1 3–9 <0.6–399

River Calder, downstream of sewage plant

24 46 18 17 3190

River Ribble 1.2 1.7 <0.38 4.4 111

River Nith <0.3–1.7 <0.6–3.5 <0.38–0.6 <0.44–2 <0.6

Avonmouth 2.4–3.6 2.9–4.7 0.6–1.0 <0.44 <0.6–7

Great Ouse and Elstow Landfill

0.4-4.2 <0.6-5.7 <0.38-1.5 <0.44-13 <0.6

River Humber 21 36 6.6 29 17

52

5.2 Biological samples5.2.1 TERRESTRIAL ECOSYSTEM

5.2.1.1 BIRDS

Muscle samples from juvenile starlings (Sturnus vulgaris) (3–4 weeks old)collected from four Swedish sites were found to contain sum levels ofBDE-47, -99 and -100 of 5.7–13 ng/g lipid (Sellström et al., 1993a; Sell-ström, 1996). The congener pattern was similar to that of Bromkal 70-5DE.No geographical trends were apparent.

5.2.1.2 Mammals PBDE levels (sum of BDEs -47, -99 and -100) were determined in rabbitmuscle (Oryctolagus cuniculus), moose muscle (Alces alces) and reindeer(Rangifer tarandus) suet samples collected within the Swedish Environ-mental Monitoring Program. PBDE were not detected in rabbit, and levelsin moose and reindeer were low, 1.7 and 0.47 ng/g lipid weight respective-ly (Jansson et al., 1993; Sellström et al. 1993a; Sellström 1996). Again, thecongener pattern was similar to that of Bromkal 70-5DE. These levels aresomewhat lower than those found in cow’s milk from Germany (4 sam-ples), which ranged from 2.5–4.5 ng/g fat measured as Bromkal 70-5DE(Krüger, 1988).

5.2.1.3 Humans Adipose tissuePreviously, DeBDE, as well as hexa-nonaBDE have been found in humanadipose tissue samples from the USA (Cramer et al., 1990; Stanley et al.,1991). The levels ranged from not detected to 1 ng/g fat for HxBDE,0.001–2 ng/g fat for HpBDE, and not detected to 8 ng/g fat for OcBDE.NoBDE levels were estimated to exceed 1 ng/g fat and DeBDE levels wereestimated to range between not detected and 0.7 ng/g fat. BDE-47 concen-tration in the adipose tissue of a Swedish 74-year old male was found to be8.8 ng/g lipid weight (Haglund et al., 1997). Human adipose tissue samplesfrom 77 individuals in Sweden collected between 1995 and 1997 were an-alyzed for 2,2',4,4'-TeBDE (Lindström et al., 1998). The means rangedfrom 3.8 to 16 ng/g lipid.

Adipose and liver tissue from two Swedish males were analyzed for sev-eral PBDEs (BDE-28, -47, -85, -99, -100, -153 and -154) (Meironyté Gu-

53

venius and Norén, 1999). The congener patterns in the two tissue types foreach individual were similar. BDE-47, -99 and -153 were the predominantcongeners with adipose BDE-47 concentrations of 2–2.4 ng/g lipid weight,BDE-99 concentrations of 1.6 ng/g lipid weight, BDE-100 concentrationsof 0.1 ng/g lipid weight and BDE-153 concentrations of 1–1.3 ng/g lipidweight. The sum of the seven PBDEs in adipose tissue was 5 ng/g lipidweight and for liver, 6 and 14 ng/g lipid weight.

Adipose tissue samples from ten randomly selected individuals in Fin-land were analyzed for BDE-47, -99 and -153 (Strandman et al., 1999).Mean concentrations were 7.3 ng/g fat for BDE-47, 2.2 ng/g fat for BDE-99 and 2.3 ng/g fat for BDE-153.

BloodA number of brominated substances have been determined in 40 humanblood plasma samples from Sweden. All samples were found to containboth TBBPA and PBDEs (Klasson Wehler et al., 1997). TBBPA levelswere in the low ng/g range on a lipid weight basis based on semiquantita-tive analyses. Six PBDE congeners (tetra-hexa BDEs) were identified andquantified: BDEs -28, -47, -66, -99, -100 and -153. Highest concentrationswere found for BDEs -47 and -99 and these made up 70% of the totalPBDE concentration in plasma. The mean concentrations of PBDEs were2.1±1.4 ng/g lipid weight.

In a study of workers at a computer disassembly plant, workers in a com-puterized office and cleaners (control), BDE-47, -153, -154, -183(2,2',3,4,4',5',6-HpBDE) as well as BDE-209 (DeBDE) were found inblood plasma for all 3 groups (Sjödin et al., 1999a). The median concen-trations (sum of 5 congeners) were highest in the computer disassemblyplant workers (26 ng/g lipid weight or 37 pmol/g lipid weight), next highestin the office workers (4.1 ng/g lipid weight or 7.1 pmol/g lipid weight) andlowest in the cleaners (3.3 ng/g lipid weight or 5.4 pmol/g lipid weight).The congener pattern was similar in the cleaners and the office workers,with BDE-47 as the dominant congener. However, the computer disassem-bly plant workers had highest median levels of BDE-183 (11 pmol/g lipidweight), followed by BDE-153 (7 pmol/g lipid weight), BDE-47 (5.9pmol/g lipid weight) and BDE-209 (5 pmol/g lipid weight). BDE-47 wasalso determined in blood serum from persons with high fish intake or nofish intake to study the influence of diet on concentrations (Bergman et al.,

54

1999). The high fish intake group had median BDE-47 concentrations of4.4 pmol/g lipid weight (2.1 ng/g lipid weight), whereas the no fish intakegroup had median concentrations of 0.83 pmol/g lipid weight (0.40 ng/g li-pid weight).

Breast milkPBDE levels in human breast milk have been determined in 25 Germanmothers (Krüger, 1988). The levels ranged from 0.6–11 ng/g fat. In a recentstudy, Norén and Meironyté, (1998; 2000) have performed a temporaltrend study of PBDE in pooled breast milk samples from Swedish mothersin Stockholm (see 6.2.4). The PBDE level (sum of 8 congeners) was 4 ng/g lipid in the 1997 sample.

PBDE levels were studied in breast milk obtained from primiparousmothers (n=39; 22-36 years old) from Uppsala county, Sweden (Darnerudet al., 1998). The individual PBDE levels found in the breast milk weresums of the five most frequently found PBDE congeners (BDEs -47, -99,-100, -153, and -154). The women also answered a questionnaire focusingon the present pregnancy, including symptoms, dietary and other habits(including smoking and alcohol consumption). Regression analysis wasused to describe a possible relationship between PBDE levels in milk andsome selected parameters from the questionnaire answers.

The observed mean value of the sum of eight PBDE congeners (sPBDE)was 4.4 ng/g fat whereas the median was 3.4 ng/g fat, which in part couldbe a consequence of a single, high peak value of 28.2 ng/g in the breastmilk from one of the women. BDE-47 was the major congener in the breastmilk, comprising ca 55% of sPBDE. Significant relationships were foundbetween milk fat levels of sPBDE and smoking (p=0.001), and betweenmilk fat levels of sPBDE and body mass index (p=0.014). However, thepresent study found no correlations between PBDE levels and the mother’sage, computer usage frequency, consumption of fish (total or specificallyfatty Baltic Sea fish), consumption of alcohol, place of residence during themother’s own childhood and adolescence (in a fishing village or not), orthe birth weight of the child. However, the number of observations in thisstudy may have been too few to reveal significant changes regarding thesecorrelations.

55

5.2.2 FRESHWATER ECOSYSTEM

5.2.2.1 FISH

Levels of BDEs -47, -99 and -100 were determined in whitefish (Coregon-us spp.) from Lake Storvindeln (pristine mountain lake in northern Swe-den), Arctic char (Salvelinus alpinus) from Lake Vättern (heavily populat-ed lake in south-central Sweden with numerous municipal and industrialpoint sources) and in trout (Salmo trutta) and pike (Esox lucius) fromseveral sites along Dalslands Canal in west central Sweden (Jansson et al.,1993; Sellström et al., 1993; Sellström, 1996). Samples were collectedbetween 1986-88 and none of these sites have known point sources forPBDE. The whitefish sample contained the lowest levels, 26 ng/g lipidweight whereas the Arctic char sample contained 520 ng/g lipid weight.BDE-47 was the predominant congener in both samples. The PBDE levelsin pike on Dalslands Canal ranged from 180–210 ng/g lipid and in troutthe range was 280–1200 ng/g lipid weight. The congener pattern was moresimilar to Bromkal 70–5DE, with similar amounts of both BDE-47 and-99. The levels are of the same order of magnitude as in the Arctic char in-dicating spread by diffuse sources. Levels of PBDE (sum of BDEs -47, -99and -100) in pike from Lake Bolmen (see section 6.2.2) for the years 1987-88 were 85–170 ng/g lipid weight, comparable to those from Dalslands Ca-nal and the congener pattern was similar as well (Kierkegaard et al. 1993).

In 1979 and 1980, high levels of TrBDEs to HxBDEs (950–27 000 ng/glipid) were found in fish sampled along the Swedish River Viskan, wherenumerous textile industries are located (Andersson and Blomkvist, 1981).BDE-47 dominated the congener pattern (70–80% of total PBDE).Theseindustries have used various brominated flame retardants in production oftextiles. No PBDEs were found in fish caught at the same sites in 1977. Thehigh levels of BDEs -47, -99 and -100 found were later confirmed in astudy where fish caught from approximately the same locations as sampledin 1987 were analyzed (Sellström et al., 1993a; Sellström, 1996). BDE-47was the predominant congener (65–96% of total PBDE). Several differentfish species were collected (pike, perch, bream, eel, tench, sea trout) inthese studies. The large differences in PBDE concentrations that werefound made it impossible to rule out species-specific differences in accu-mulation, thus making it difficult to draw conclusions about the location ofthe sources along the river.

56

New samples of pike as well as sediments (see section 5.1.4) were col-lected along the River Viskan (Figure 7) in 1995 in order to search forsources. Pike samples were obtained at only 4 of the 8 sites. BDEs -47,-99, -100 and -209 (DeBDE) as well as HBCD were quantified in the fish.BDE-209 was found in trace amounts in a few fish. The concentrations ofthe other substances increased further downstream as more industries werepassed (Sellström et al., 1998b). The concentrations of BDEs -47, -99 and-100 together ranged from not detected to 4600 ng/g lipid weight, withBDE-47 again being the predominant congener (50–90%), and HBCDranged from not detected to 8000 ng/g lipid weight (Figure 10). The lowestlevels of the PBDEs and HBCD were found upstream of the industries.This is the first time HBCD has been found in environmental samples inSweden.

Loganathan et al. (1995) found TeBDE to HxBDE in carp (Cyprinuscarpio) from Buffalo River, New York, an area around the Great Lakesshowing environmental impairment. TeBDE dominated the congener pat-tern (94–96% of total PBDE) and TeBDE and PeBDE levels were 13–22ng/g fresh weight. Recently, Asplund et al. (1999b) found tri- to hexaBDEsin steelhead trout (Oncorhynchus mykiss) from Lake Michigan sampled in1995. The sum of BDE-47, -99 and -100 was 2700 ng/g lipid weight. Laketrout (Salvelinus namaycush) from several Great Lakes were also found to

0

100

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BDE47 in Pike

No fishcaught

No fishcaught

Sampling Location Sampling Location

BDE209 and HBCD in Pike

ng/gl.w.

ng/gl.w.

Figure 10. BDE-47, BDE-209 and HBCD concentrations in pike from RiverViskan (Sellström et al., 1998b).

57

have di- to heptaBDEs with sum concentrations of 540 ng/g lipid weightfor Lake Ontario, 240 ng/g lipid weight for Lake Huron and 140 ng/g lipidweight for Lake Superior (Alaee et al., 1999).

In eels (Anguilla anguilla) from Dutch rivers and lakes (10 locations),levels of BDE-47 ranged from <20 to 1700 ng/g lipid and BDE-47 com-prised 70% of the total PBDE (de Boer, 1990). Several species of freshwa-ter fish from waters of North-Rhine Westphalia contained 18–983 ngPBDE/g lipid (Krüger, 1988).

Allchin et al. (1999) have recently carried out a survey of PBDE in sed-iments (see section 5.1.4) from several United Kingdom rivers and estuar-ies and fish from the estuaries. Plaice (Pleuronectes platessa), flounder(Platichthys flesus) and dab (Limanda limanda) were collected in the estu-aries of rivers with suspected sources including a manufacturer of PeBDEand OcBDE, several industries using PeBDE, several landfills receivingwastes suspected to contain PBDEs and a reference site (Table 10). The re-sults are given in Table 12, and these support the conclusions drawn previ-ously for sediments – that a major source is the manufacturing plant on theRiver Tees. The predominant congener in fish is BDE-47, particularlywhere sediments are highly contaminated. This is similar to the situationfound in Sweden along the River Viskan. It is also interesting to note thatOcBDE as the technical product DE-79, is bioavailable and found in rela-tively high concentrations in fish exposed via sediments.

Table 12. Concentrations of BDE-47, -99, PeBDE quantified as the technical product DE-71 and OcBDE quantified as the technical product DE-79 in fish from UK river estuaries(ng/g lipid weight). No DeBDE quantified as the technical product DE-83 was detected inany samples. Data from Allchin et al. 1999.

Site Species BDE-47 BDE-99 DE-71 (PeBDE)

DE-79 (OcBDE)

Tees Bay Plaice, flo-under, dab

520–9500 83–370 920–1200 500–1200

Lune/Wyre, Off River Calder

Flounder 400 54 100 120

Nith estuary Flounder 73-120 nd-19 47–120 nd-83

Bideford Bay(Off Avonmouth)

Flounder, plaice, dab

nd-370 nd-100 94–120 nd-970

The Wash (Off Great Ouse)

Dab 380 74 110 58

58

5.2.2.2 BirdsMuscle samples from ospreys (Pandion haliaetus) found dead in variousparts of Sweden were pooled and analyzed for PBDE (Jansson et al., 1993;Sellström et al., 1993a; Sellström, 1996). Ospreys feed on freshwater fish.The PBDE concentration was 2100 ng/g lipid (sum of BDEs 47, 99 and100) with BDE-47 dominating the congener pattern (86%). These high lev-els may reflect biomagnification and/or reflect fish consumption alongtheir migratory routes to Africa.

5.2.3 MARINE ECOSYSTEM

5.2.3.1 FISH AND SHELLFISH Hepatopancreas samples from Dungeness crab from the several sites on theStrait of Georgia, British Columbia, Canada, were analyzed for di- to hep-taBDEs (Ikonomou et al., 1999). BDE-47 was the major congener foundand the sums of BDE-47 and -99 were approximately 100–350 ng/g lipidweight.

The sum of BDE-47, -99 and -100 in fall-caught herring (Clupea haren-gus) muscle from five sites along the Swedish coast ranged from 17-62 ng/g lipid, with BDE-47 being the dominant congener (Sellström et al., 1993a;Sellström, 1996). Similarly, BDE-47 levels in different age groups of Bal-tic herring ranged from 3.2 to 27 ng/g lipid with the sum of BDE-47, -99and -100 ranging from 3.2 to 32 ng/g lipid (Haglund et al., 1997). Lowestlevels were in 2-year-old herring and highest levels were in 5-year-old her-ring. Haglund et al. (1997) found a similar trend for methoxy-PBDE intheir Baltic herring samples. Strandman et al. (1999) also found increasingconcentrations of BDE-47, -99 and -153 with age in Baltic sprat (age 3–13years) (Sprattus sprattus) but not in herring. BDE-47 was the major conge-ner found and concentrations ranged from 7.6-24 ng/g lipid weight for 1-to 3-year-old sprat, 17–140 ng/g lipid weight for 3- to 13-year-old sprat and7.6–24 ng/g lipid weight in the herring. Whole body composites of herringwere found to have BDE-47, -99 and -100 concentrations of 6.2, 0.6 and0.8 ng/g lipid and sprat had 4.3, 0.7 and 0.8 ng/g lipid (Burreau et al.,

Off River Humber Flounder 1600 160 110 900

Site Species BDE-47 BDE-99 DE-71 (PeBDE)

DE-79 (OcBDE)

59

1999). The levels found in Baltic herring are similar to BDE-47 levels of8.4-100 ng/g lipid found by de Boer (1990, 1995) in herring from three re-gions in the North Sea.

BDE-47, -99 and -153 levels in Baltic salmon (Salmo salar) musclewere 167, 52 and 4.2 ng/g lipid, repectively (Haglund et al., 1997). Meth-oxy-PBDE were also found. In whole body composites, BDE-47, -99 and-100 levels were 47, 7.2 and 6.3 ng/g lipid (Burreau et al., 1999). In anotherstudy, muscle, ripe eggs and blood plasma from Baltic salmon were ana-lyzed for a range of organohalogen compounds including BDE-47, -99 and-100 (Asplund et al., 1999a). The levels found are shown in Table 13. Sev-eral hydroxylated and methoxy-PBDE were also found. Methoxy-PBDEwere found in all samples at similar concentrations to the PBDEs. Severalhydroxylated PBDEs were found in blood samples at 20–30% of the meth-oxy-PBDE levels.

Table 13. Mean concentrations of PBDEs in tissues from Baltic salmon (ng/g lipid weight).Results from Asplund et al. (1997; 1999a).

Cod (Gadus morhua) liver collected from three regions of the North Seahad sum levels of BDEs -47 and -99 of 1.9–360 ng/g lipid (de Boer, 1989;1995). Watanabe et al. (1987) found PBDE in several marine fish andshellfish samples in Japan. TeBDE and PeBDE concentrations were be-tween 0.1 and 17 ng/g fresh weight with TeBDE being the major compo-nent in the samples. A mussel sample from Osaka Bay was also found tocontain DeBDE.

5.2.3.2 Birds Previously, Di- and TrBDE have been identified in black skimmer tissuesand eggs (Rynchops nigra) in the U.S. but no quantitative analysis could beperformed due to lack of standards (Stafford, 1983). PBDE levels weremeasured in white-tailed sea eagle collected from the Baltic Sea and foundto contain 350 ng/g lipid weight (Jansson et al., 1987). Common guillemots(Uria aalge) collected in 1979–81 from the Baltic and North Seas con-

Tissue BDE-47 BDE-99 BDE-100

Muscle 190 52 46

Ripe eggs 64 16 18

Blood 190 55 59

60

tained 370 and 80 ng PBDE/g lipid (Jansson et al., 1987). Brunnich’s guil-lemot (Uria lomvia) from Svalbard in the Arctic contained 130 ng PBDE/g lipid) (Jansson et al., 1987). More recent results of PBDE analyses inguillemot eggs from the Baltic Sea are given in section 6.2.3.

In Sweden, PBBs have been analyzed in a number of samples from ani-mals in different ecosystems (Jansson et al., 1993) including the Arctic(Jansson et al., 1987). The levels seen are quite low.

5.2.3.3 Mammals Several species of seal from several different sites have been analyzed forPBDE. Female grey seals (Halichoerus grypus) from the Baltic Sea col-lected in 1979–1985 contained 730 ng PBDE/g lipid in their blubber (sumof BDEs -47, -99 and -100) (Jansson et al., 1993; Sellström et al., 1993a;Sellström, 1996). Male Baltic grey seals had 280 ng/g lipid weight of sumBDEs -47, -99 and -100 and male ringed seals (Pusa hispida) had 320 ng/g lipid weight (Andersson and Wartanian, 1992). Blubber from Baltic greyand ringed seals collected between 1981 and 1988 were found to contain419 ng PBDE/ g lipid and 350 ng PBDE/g lipid (sum of BDEs 47, 99 and100), respectively (Haglund et al., 1997). Methoxy PBDE were alsopresent in both species.

Female ringed seals collected in 1981 from Svalbard in the Arctic con-tained 40-51 ng PBDE/g lipid (Jansson et al., 1987; 1993; Sellström et al.,1993a; Sellström, 1996). Ringed seal from the Canadian Arctic had meanPBDE concentrations (di- to hexaBDEs) of 25.8 ng/g lipid weight (fe-males) and 50 ng/g lipid weight (males) (Alaee et al., 1999). Harbour seal(Phoca vitulina) from the Baltic Sea contained 90 ng PBDE/g lipid as com-pared to harbour seal from the North Sea, which contained 10 ng PBDE/glipid (Jansson et al., 1987; Sellström, 1996). Andersson and Wartanian(1992) found 230 ng PBDE/g lipid in harbour seals from the Swedish westcoast (Skagerrak).

Blubber samples from three bottlenose dolphins (Tursiops truncatus)collected during a mass mortality event on the south Atlantic U.S. coast in1987/88 contained 180-220 ng PBDE/g lipid (Kuehl et al., 1991). Bottle-nose dolphins from the Gulf of Mexico were found to contain up to 8000ng PBDE/g lipid (Kuehl and Haebler, 1995). De Boer et al. (1998a; 1998b)have found PBDE and PBB in the blubber of three sperm whales (Physetermacrocephalus), one minke whale (Balaenaoptera acutorostrata) and one

61

whitebeaked dolphin (Lagenorhynchus albirostris) found stranded on theDutch coast in early 1998. Several harbour seals were also sampled. Spermwhales feed in deep water and the stranded whales’ stomachs were empty,indicating that the exposure occurred in the deep Atlantic via the food web.The levels found are given in Table 14. Analyses included BDE-209(DeBDE) but levels were below detection limits in all samples. Long-finned pilot whale (Globicephala melas) from the Faeroe Islands in thenorth Atlantic were analyzed for 19 PBDEs (Lindström et al., 1999). High-est concentrations were found in young males and females (3000–3160 ng/g lipid) compared to adult females (840–1050 ng/g lipid) and males (1610ng/g lipid). In a second study of long-finned pilot whales, a similar trendwas seen with young animals having PBDE concentrations of 740 ng/g li-pid weight, adult females having 230 ng/g lipid and adult males having 540ng/g lipid (van Bavel et al., 1999).

Beluga (Delphinapterus leuca) from the Canadian Arctic had meanPBDE concentrations (di- to hexaBDEs) of 81.2 ng/g lipid weight (fe-males) and 160 ng/g lipid weight (males) (Alaee et al., 1999). The sum ofBDE-47 and -99 concentrations in a killer whale (Orca orcinus) and sev-eral porpoises from British Columbia, Canada were 100 ng/g lipid weightand 300–2000 ng/g lipid weight, respectively (Ikonomou et al., 1999).

The congener profile in marine mammals shows highest concentrationsfor BDE-47.

Table 14. Levels of PBDE in blubber (ng/g lipid weight) from several marine mammalscollected along the coast of the Netherlands. Results from de Boer et al. (1998b).

Species Fat content in percent

BDE-47 BDE-99 BDE-100

Sperm whale 1 72.2 130 36 21

Sperm whale 2 23.4 250 64 35

Sperm whale 3 31.7 190 32 24

Minke whale 14 630 160 79

White-beaked dolphin 99 5500 1000 1200

Harbour seal 1 24.4 4900 660 450

Harbour seal 2 96.3 1250 42 100

Harbour seal 3 72.2 390 190 25

62

5.3 Bioaccumulation, biomagnificationBioconcentration factors from water to Baltic blue mussels were found tobe 1 300 000 for BDE-47, 1 400 000 for BDE-99 and 220 000 for BDE-153 (Gustafsson et al., 1999). These were higher than for the tri-hexaCBcongeners that were also tested.

When sediment concentrations were compared to those in pike collectedat several of the same sites along the River Viskan (see sections 5.1.4 and5.2.2.1), high fish to sediment ratios were seen for BDE-47, -99, -100 andHBCD indicating that these are highly bioavailable (Table 15).

Table 15. Ratios between the concentrations of PBDE and HBCD found in fish (ng/g lipid)and the concentrations found in sediments (ng/g carbon) (Sellström et al., 1998b).

* – means concentrations in fish and/or sediment were below detection limits

Bioaccumulation has been studied in zebrafish fed with dried chironimidlarvae treated with BDEs -28, -47, -66, -85, -99, -100, -138, -153 and -154(Andersson et al., 1999). Highest accumulation was seen for BDE-47 fol-lowed by BDE-28. BDE-100, -153 and -154 accumulated to a lesser extent.BDE-99 did not accumulate, which is not in agreement with the results ofGustafsson et al. (1999) in blue mussels or with uptake efficiency data seenin pike by Burreau et al. (1997).

Concentrations of PBDE in herring and their predators grey seal andguillemot, all collected in the same area of the Baltic Sea, have been com-pared to estimate potential biomagnification (Sellström, 1996). The her-ring were caught in the autumn of the same year as guillemot egg collection(1987). The grey seal sample was a pooled sample from 8 females founddead in the area during 1979–1985. In a recent study, herring, sprat andsalmon were collected from the Baltic Proper in 1998 (Burreau et al.,1999). These were analyzed for PBDEs (BDEs -17+25, -28, -35, -47, -49,-66, -99, -100 and -154) and nitrogen isotopes to study biomagnification.

Location BDE-47 BDE-99 BDE-100 BDE-209 HBCD

1 –* – – – –

2 19 17 36 – –

3 17 – 11 – 15

4 6.6 – 4.6 – 0.6

9 – – 30 – –

63

Salmon feed primarily on sprat and the lipid weight concentrations in thetwo species can be compared. The calculated biomagnification factors aregiven in Table 16.

BDE-47 appears to biomagnify to the largest extent, which is in agree-ment with the results of uptake studies in fish showing highest assimilationefficiences for BDE-47 from the gut compared to BDE-99 and -153 (Bur-reau et al., 1997). Burreau et al. (1999) also calculated b values (biomag-nification potential) for all BDE congeners studied and found that thesewere all positive, meaning that all the studied congeners biomagnify. How-ever, there were differences, with tetra- and pentaBDEs biomagnifying toa similar degree, the triBDEs biomagnifying somewhat less and the hex-aBDE biomagnifying considerably less.

Table 16. Biomagnification factors of PBDE congeners in the guillemot and grey seal (ng/g lipid weight) compared to herring (ng/g lipid weight) from the same area of the Baltic Sea(Sellström, 1996) and for Baltic salmon compared to sprat (Burreau et al., 1999).

Osprey have higher concentrations than most of the freshwater fish an-alyzed in Swedish studies but no firm conclusions about biomagnificationcan be drawn due to their migratory habits.

Species BDE-47 BDE-99 BDE-100

Guillemot egg/Herring 19 17 7.1

Grey seal/Herring 19 4.3 6.8

Salmon/Sprat 11 10 8

64 Table 17. An overview of concentrations of several PBDEs, HBCD and TBBPA in environmental samples: air in pg/m3; sewage sludge in ng/g dryweight; sediments in ng/g ignition loss; biota in ng/g lipid weight (unless otherwise stated). The sum of PBDE is the sum of BDE-47, -99 and -100but, if more congeners are included, this is marked with an asterisk (*).

Sample type Location BDE-47 BDE-99 BDE-100 Sum of PBDE BDE-209 HBCD TBBPA Reference

Air Ammarnäs 6.3 1.6 0.4 8.3 nd 6.1 1

Air Hoburgen 0.7 0.35 0.07 1.1 nd 5.3 1

Air Stoke Ferry, UK 4.7-50 5.5-13 1.1-3.9 11-67 33

Air Hazelrigg, UK 3.2-61 3.1-22 0.62-5.4 6.9-28 33

Air Alert, Canada 1-28* 34

Air Dunai Is., Russia 1-8* 34

Sewage sludge Gothenburg 15 19 3.5 38 2

Sewage sludge Klippan 22 18 5.4 45.4 56 7, 35

Sewage sludge Rimbo 53 53 13 119 31 7, 35

Sewage sludge 3 plants, Stockholm 39-91 48-120 11-28 98-239 140-350 11-120 2.9-76 29

Sewage sludge Germany 0.4-15* 3

Sediment Japan 21-59 (dw) <25-11600 (dw)

4, 5

Sediment Baltic Sea nd-3.4 nd-2.4 nd-1.3 nd-5.4 2, 6

Sediment Upstream plastics plant

3.7 8.8 1.6 14.1 50 7

Sediment Downstream plastics plant

780 1200 270 2250 430 7

65

Sediment R. Viskan, upstream and downstream textile industries

<2-50 <1-53 <0.4-19 nd-120 nd-16000 nd-7600 8

Sediment 22 European river mouths

<0.17-6.2 (dw)

<0.19-7.0 (dw)

<0.51-1800 (dw)

32

Sediment 7 rivers, Great Britain

<0.3-368 (dw)

<0.6-898 (dw)

<0.6-3190 (dw)

28

Terrestrial ecosystem

Starlings (young) Sweden 2.7-7.8 2.3-4.2 0.6-1.1 5.7-13 9, 10

Rabbit Sweden <1.8 <0.3 <0.2 <2.3 9, 10

Moose Sweden 0.8 0.6 0.2 1.7 9, 10

Reindeer Sweden 0.17 0.26 0.04 0.47 9, 10

Cow's milk Germany 2.5-4.5* 11

Human blood Sweden 2.1* present ng/g

12

Human blood Sweden – computer disassembly work-ers

2.9 (median)

26* 4.8 31

Human blood Sweden – cleaning personnel/Office workers

1.5-1.6 (median)

3.3-4.1* <0.7 (median)

31

Human blood Sweden – high fish intake

2.1 38

Sample type Location BDE-47 BDE-99 BDE-100 Sum of PBDE BDE-209 HBCD TBBPA Reference

66

Human blood Sweden – no fish intake

0.40 38

Human adipose Sweden 8.8 1.1 1.8 11.7 13

Human adipose Sweden 3.8-16 14

Human adipose Sweden 2.2 1.6 0.1 5* 36

Human adipose USA nd-0.7 15

Human adipose Finland 7.3 2.3 6.2-22* 37

Human breast milk

Germany 0.6-11* 11

Human breast milk

Sweden 2.3 0.5 0.4 4* 16

Human breast milk

Sweden 2.5 0.7 0.5 4.4* 17

Freshwater ecosystem

Whitefish L. Storvindeln 15 7.2 3.9 26 9, 10

Arctic char L. Vättern 400 64 51 520 9, 10

Pike L. Bolmen, 1993 65 42 19 130 18

Pike Dalslands canal 94-98 60-79 25-36 180-210 9, 10

Pike R. Viskan, upstream and downstream

<46-2000 <37-1600 <14-1000 <130-4600 trace <50-8000 8

Trout Dalslands canal 120-460 130-590 33-150 280-1200 9, 10

Sample type Location BDE-47 BDE-99 BDE-100 Sum of PBDE BDE-209 HBCD TBBPA Reference

67

Carp USA 13-22* (fw) 19

Steelhead trout Lake Michigan, USA

1700 600 360 3000* 39

Lake trout Lake Ontario, USA 540* 34

Lake trout Lake Huron, USA 240* 34

Lake trout Lake Superior, USA 140* 34

Eels Netherlands <20-1400 <50-1700 20

Several fish species

Germany 18-983* 11

Osprey Sweden 1800 140 200 2140 9, 10

Marine ecosystem

Herring Baltic Sea 19-38 7.8-17 3.4-6 30-61 9, 10

Herring Baltic Sea 3.2-27 nd-2.9 1.3-1.9 3.2-32 13

Herring Baltic Sea 7.6-24 4.3-3.9 12.9-28.3* 37

Herring Baltic Sea 6.3 0.6 0.8 12* 43

Herring Kattegatt 12 3.4 1.6 17 9, 10

Herring North Sea 8.4-100 20, 21

Sprat (different age groups)

Baltic Sea 17.5-140.8 1.9-9.5 21-149* 37

Sprat Baltic Sea 4.3 0.7 0.8 8.4* 43

Cod liver North Sea 170 1.9-360 21, 22

Sample type Location BDE-47 BDE-99 BDE-100 Sum of PBDE BDE-209 HBCD TBBPA Reference

68

Salmon Baltic Sea 167 52 44 220 13

Salmon Baltic Sea 190 52 46 290 23

Salmon Baltic Sea 46 7.3 6.4 86* 43

Several fish species

Japan 0.1-17* (fw) 4

Several flatfish species

7 river estuaries, Great Britain

73-9500 16-790 nd 28

White-tailed sea eagle

Baltic Sea 350 24

Common guillemot

Baltic Sea, 1994 480 36 61 570 130 10, 30

Brunnich's guillemot

Svalbard 130 24

Cormorant liver Rhine delta 28000 (fw) 20

Grey seal Baltic Sea 650 40 38 730 9, 10

Grey seal Baltic Sea 308 54 57 419 13

Grey seal Baltic Sea 280 40

Ringed seal Baltic Sea 256 33 61 350 13

Ringed seal Baltic Sea 320 40

Ringed seal Svalbard 47 1.7 2.3 51 9, 10

Ringed seal Canadian Arctic 25.8-50* 34

Harbour seal Baltic Sea 90 24

Sample type Location BDE-47 BDE-99 BDE-100 Sum of PBDE BDE-209 HBCD TBBPA Reference

69

nd - not detected. dw - dry weight. fw - fresh weight.

References: 1. Bergander et al. (1995); 2. Nylund et al. (1992); 3. Hagenmeier et al. (1992); 4. Watanabe et al. (1986; 1987; 1995); 5. Environmental Agency Japan(1991); 6. Jonsson and Kankaanpää (1999); 7. Sellström and Jansson (1995); 8. Sellström et al. (1998b); 9. Sellström et al. (1993a); 10. Sellström (1996); 11.Krüger (1988); 12. Klasson Wehler et al. (1997); 13. Haglund et al. (1997); 14. Lindström et al. (1998); 15. Cramer et al. (1990); Stanley et al. (1991); 16. Norénand Meironyté (1998; 2000); 17. Darnerud et al. (1998); 18. Kierkegaard et al. (1993); 19. Loganathan et al. (1995); 20. de Boer (1990); 21. de Boer (1995); 22. deBoer (1989); 23. Asplund et al. (1999a); 24. Jansson et al. (1987); 25. de Boer et al. (1998b); 26. Kuehl et al. (1991); 27. Kuehl and Haebler (1995); 28. Allchin etal. (1999); 29. Sellström et al. (1999); 30. Kierkegaard et al. (1999b); 31. Sjödin et al. (1999a); 32. van Zeil (1997); 33. Peters et al. (1999b) and A. Kierkegaard,personal communication; 34. Alaee et al. (1999); 35. Sellström (1999); 36. Meironyté Guvenius and Norén (1999); 37. Strandman et al. (1999); 38. Bergman et al.(1999); 39. Asplund et al. (1999b); 40. Andersson and Wartanian (1992); 41. Lindström et al. (1999); 42. van Bavel et al. (1999); 43. Burreau et al. (1999).

Harbour seal Skagerrak 230 40

Harbour seal North Sea 390-4900 42-660 25-450 600-6000 25

Bottlenose dolphin

S. Atlantic 180-220 26

Bottlenose dolphin

Gulf of Mexico 8000 27

White-beaked dolphin

Netherlands 5500 1000 1200 7700 25

Beluga Canadian Arctic 81-160* 34

Long-finned pilot whale

Faeroe Islands 410-1780 160-600 87-280 843-3160* 41

Long-finned pilot whale

Faeroe Islands 66-860 24-170 12-98 126-1250* 42

Minke whale Netherlands 630 160 79 870 25

Sperm whale Netherlands 130-250 32-64 21-35 187-349 25

Sample type Location BDE-47 BDE-99 BDE-100 Sum of PBDE BDE-209 HBCD TBBPA Reference

70

6. Trends

6.1 Spatial trendsResults of PBDE analyses in surficial sediments from the Baltic Sea andthe mouths of major European rivers indicates a gradient, with highest con-centrations in southern Europe and lower levels in Scandinavia and theBaltic Sea (Kierkegaard et al., 1997; Sellström et al., 1999).

Results of PBDE analyses in freshwater fish in Sweden indicate thatsouthern Sweden is more contaminated with PBDE than northern Sweden(Sellström 1996).

Fall caught herring collected at five sites along the Swedish coast showthat the lowest concentrations are found on the west coast (17 ng/g lipid)and the highest are found in the southern part of the Baltic Sea (62 ng/glipid). The concentrations in the Baltic Sea then decrease from south tonorth up to Bothnian Bay (30 ng/g lipid) (Sellström et al., 1993a; Sell-ström, 1996). This spatial trend is almost identical to that found previouslyfor PCB and DDT (NV, 1988).

De Boer (1989, 1995) found a clear spatial trend for PBDE in cod liverin the North Sea, with decreasing levels from south (22–360 ng/g lipid) tonorth (1.9–68 ng/g lipid). This was attributed to major inputs from rivers inwestern Europe. A similar trend was seen for herring although the trendwas not as clear due to the migration patterns of the herrring (de Boer,1990).

A study of common guillemots collected in 1979–81 from the Baltic andNorth Seas and Brunnich’s guillemot from Svalbard in the Arctic, indicat-ed higher PBDE levels in the Baltic Sea (370 ng/g lipid) compared to theArctic (130 ng/g lipid) (Jansson et al., 1987). Cormorant liver from theRhine River delta have very high levels (28000 ng/g fresh weight) (de Bo-er, 1990). Similarly, harbour seal from the Baltic Sea had higher levels (90ng/g lipid) than in the northern North Sea (10 ng/g lipid) but highest levelswere seen along the coast of the Netherlands (600–6000 ng/g lipid) (deBoer et al., 1998b). Low PBDE levels are generally seen in Arctic ringedseals (26–51 ng/g lipid) (Jansson et al., 1987; Alaee et al., 1999) comparedto the Baltic Sea (320–350 ng/g lipid) (Haglund et al., 1997; Anderssonand Wartanian, 1992).

71

Together, these results indicate that southern European coastal areas aremost contaminated by PBDE, followed by the Baltic Sea. A gradient is alsoindicated in northern Europe with levels declining from south to north,with lowest levels in the Arctic. These results are similar to those seen fororganochlorine contaminants such as PCB and DDT.

6.2 Temporal trends6.2.1 BALTIC SEA SEDIMENT CORE

A laminated sediment core was collected from the southern part of the Bal-tic Proper for analysis of a number of organochlorine contaminants as wellas PBDEs (Nylund et al., 1992). The core was dated by counting thenumber of laminae and assuming that each laminae represents one year’sdeposition. The core was divided into sections and each section analyzed.The results provide a retrospective time trend from 1939 to 1987 (Figure11) and show that the PBDE levels (sum of BDEs -47, -99 and -100) haveincreased, particularly after 1980. PBDE level in the sample from 1989was 2.9 ng/g ignition loss.

0

0,5

1

1,5

2

2,5

3

3,5

1939

1946

1953

1961

1967

1971

1974

1976

1978

1980

1982

1984

1986

1987

Year

ng

/g ig

nit

ion

loss

Figure 11. Concentrations of PBDE in different layers of a sediment corerepresenting the years 1939–1987 (Nylund et al., 1992).

72

6.2.2 PIKE FROM LAKE BOLMEN

Pike are collected yearly from Lake Bolmen within the National Environ-mental Monitoring Program for contaminant analyses. Samples are alsobanked at the Enviromental Specimens Bank at the Swedish Natural His-tory Museum in Stockholm. A retrospective time trend for PBDEs was de-termined by analyzing BDEs -47, -99 and -100 in pooled samples for mostyears between 1967 and 1990. For the years 1974, 1981, 1987 and 1991-1996, 9–11 individual muscle samples were analyzed for each year. The re-sults show significantly increasing trends for the three congeners from1967 to the early 1980s. The trend for BDE-47 is shown in Figure 12. From1982 to 1996, there are large between-year variations, however, thereseems to be a tendency towards a fairly even trend level with no indicationof decreasing levels (Kierkegaard et al., 1993, Kierkegaard et al. 1999b).The predominant congener is BDE-47. The PBDE level (sum of BDE-47,-99 and -100) was 100 ng/g lipid weight for 1996. Based on the established

Figure 12. Concentrations of BDE-47 in Lake Bolmen pike (Kierkegaard et al.,1999b). The line represents a three-point running mean smoother (p<0.001). Cir-cles represent arithmetic means with bars indicating 95% confidence intervals.

73

time trend, yearly analysis of PBDEs in pike from Lake Bolmen is now in-cluded within the National Environmental Monitoring Program.

Methoxy-BDE-47 was also analyzed in the pike samples and the tempo-ral trend for these is shown in Figure 13. The concentrations show a signif-icantly decreasing trend. The origin of this is not known. Biogenic sourcesfrom primary producers could be one explanation, however, from 1966 to1997, this lake has had an increasing eutrophication trend (Kierkegaard etal., 1999b).

6.2.3 ROACH FROM LAKE KRANKESJÖN

A retrospective temporal trend study was also performed using roach (Ru-tilus rutilus), which are collected yearly from Lake Krankesjön, a eutro-phied lake, within the National Environmental Monitoring Program forcontaminant analyses. Samples are also banked at the Enviromental Spec-imens Bank at the Swedish Natural History Museum in Stockholm. A ret-

Figure 13. Concentrations of methoxy-BDE-47 in pike from Lake Bolmen(Kierkegaard et al., 1999b). The curve represents a log-linear regression line(p<0.001). Circles represent arithmetic means and bars indicate 95% con-fidence intervals.

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rospective time trend for PBDEs was determined by analyzing BDEs -47,-99 and -100 in individual samples for several years (1980, 1983, 1985,1988, 1990, 1992, 1994 and 1996) (Kierkegaard et al., 1999b). The con-centrations in roach were generally lower than in pike and there was con-siderable between year variation as well as within years. No significanttrend was detected (Figure 14). No methoxy-BDE-47 was detected.

6.2.4 GUILLEMOT EGGS FROM ST. KARLSÖ

Samples of guillemot eggs (St. Karlsö, Baltic Sea) are collected yearlywithin the National Environmental Monitoring Program. Samples are alsobanked at the Enviromental Specimens Bank at the Swedish Natural His-tory Museum in Stockholm. A retrospective time trend for PBDEs was de-termined by analyzing BDEs -47, -99 and -100 in pooled egg samples fromthe specimen bank for most years between 1969 and 1990 (Sellström et al.,1993a; b; Sellström, 1999). Samples from 10 individuals were analyzed for

Figure 14. Concentrations of BDE-47 in Lake Krankesjön roach (Kierkegaardet al., 1999b). Circles represent geometric means of 8 individuals/year andbars represent 95% confidence intervals.

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the years 1975, 1989 and 1992-1997. The results show significantly in-creasing trends for these three congeners from 1969 to the beginning of the1990s. The results for BDE-47 and -99 are shown in Figure 15. There islarge between-year variation after 1990, but statistical analysis indicatesthat PBDE levels have declined for the period 1992–1997. The predomi-nant congener is BDE-47 and the PBDE level (sum of BDE-47, -99 and-100) in 1997 was 190 ng/g lipid. Based on the established time trend, year-ly analysis of PBDEs in guillemot eggs from St. Karlsö is now includedwithin the National Environmental Monitoring Program.

69 74 79 84 89 94 69 74 79 84 89 94

Figure 15. Concentrations of BDE-47 and -99 in guillemot eggs collected fromSt. Karslö (Sellström et al., 1993a; b; Sellström, 1999). Circles represent thearithmetic means and bars indicate 95% confidence intervals. The line repre-sents a five-point running mean smoother (p<0.001).

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A temporal trend for HBCD has also been studied in the same guillemotsamples as were analyzed for PBDEs (Kierkegaard et al. 1999b). The trendshows increasing levels from 1969 to 1997.

6.2.5 HUMAN BREAST MILK

Pooled human milk samples collected during the period 1972 to 1997 fromnative Swedish mothers living in the Stockholm region were analyzed for8 PBDE congeners (Figure 16). All 8 PBDEs were present in most samplesand the predominant congener was BDE-47 (Meironyté et al., 1998; Mei-ronyté et al., 1999).

The results show an exponential increase of PBDEs in human breastmilk from 1972 to 1997 with a doubling rate of 5 years (Norén and Meiro-nyté, 1998; 2000) (Figure 17). The PBDE level (sum of 8 congeners) was4 ng/g lipid in the 1997 sample. This time trend differs considerably fromthose of pike and guillemot, which have leveled off or may be decreasing.This may indicate that the current exposure in humans may not be from just

0

0,5

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1,5

2

2,5

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,4, 4

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,4,4

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2,2'

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ng/g

lipi

ds

1972

1976

1980

1984/1985

1990

1994

1996

1997

Congeners of brominated diphenyl ethers

Figure 16. Brominated diphenyl ethers in pooled samples of human milk, sam-pled 1972–1997 (Meironyté et al., 1998).

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diet. Possible other routes of exposure could be the presence of brominatedflame retardants in the work and home environment.

PBDE

doubling 5 years

pg/g

lipi

d

y = 91,169e0,1395x

R2 = 0,914

PBDE sum Expon. (PBDE sum)

1972

1976

1980

1984

1990

1994

1996

4000

3500

3000

2500

2000

1500

1000

500

0

Figure 17. Concentrations of PBDE (sum of 8 congeners) in pooled humanbreast milk samples from different time periods (Norén and Meironyté, 1998;2000).

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7. Summary and conclusionsThe sources of brominated flame retardants to the environment identifiedin Sweden are plastics and textile industries.The work environment alsocontains these as shown from their presence in indoor dust and air. Thepresence of PBDE and PBB in samples from living organisms in the Arctic(Svalbard) and the presence of PBDE in air samples from southern Gotlandand northern Sweden indicate long range transport of these in air. The pres-ence of BDE-47, -99 and -209, TBBPA and HBCD in sewage sludge indi-cates sources emptying into sewage systems, either from households, traf-fic and/or diffuse releases to the environment. Lower brominated PBDEhave been found at low concentrations in sediments from the Baltic Seabut, together with DeBDE, in higher concentrations at the mouths of someEuropean rivers, in several United Kingdom rivers and in the Swedish Riv-er Viskan.

Levels of BDE-47, -99 and -100 are low in mammals and birds from theterrestrial ecosystem. Higher concentrations of these have been found inbiota samples (fish, birds, mammals) in aquatic and marine ecosystems. Aspatial trend is apparent with highest levels in biota from the Netherlandscoast followed by the Baltic Sea, with lower levels in the North Sea andlowest levels in northern Sweden and the Arctic. The spatial trend is verysimilar to that found for PCB and DDT.

BDE-47 dominates in samples from areas affected by pollution (LakeVättern, Baltic Sea, River Viskan), whereas the congener pattern is moresimilar to the technical product Bromkal 70-5DE in sewage sludge, sedi-ments and fish from background sampling sites. The highest concentra-tions of BDE-47 in Sweden are found in fish collected along the RiverViskan where several textile industries are located. Sediments from thisriver have also been found to contain DeBDE and HBCD, and fish alsocontain HBCD. High concentrations of especially BDE-47 are also foundin fish-eating birds and mammals, possibly due to bioaccumulation and bi-omagnification since this congener has the highest bioavailability. In fishfrom the United Kingdom, highest concentrations of BDE-47, -99, OcBDEand DeBDE are found in estuaries of rivers with manufacturing plants oruser sites upstream.

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Lower brominated PBDE, OcBDE and HBCD are bioavailable fromsediment, as indicated by their presence in fish. Uptake from the gastroin-testinal tract of rats, mice and fish is high for lower brominated PBDE.High rates of transfer are also seen in neonatal mice via breast milk. Spe-cies-specific differences are seen in metabolic capacity however. Metabo-lism seems to result in the formation of hydroxylated PBDE. HBCD is alsorapidly absorbed from the gut in rats. Uptake from the GI tract of fish islow for DeBDE, but metabolic debromination to lower brominated PBDEmay occur.

BDE-47, -99 and -100 biomagnify in fish-eating birds and mammals.BDE-47, -99 and -100 have been found in human and Baltic salmon blood.PBDE have been found in human adipose tissue, blood and breast milk.Higher brominated PBDE, including DeBDE, have been found in humanadipose tissue and in blood. Highest DeBDE concentrations were found inpeople working at a computer disassembly facility. DeBDE is also bioa-vailable in fish.

Several hydroxylated and methoxylated PBDE (Te- and PeBDE) havebeen found in salmon, herring, ringed seal and grey seal from the Baltic Seaand methoxy-BDE-47 in freshwater pike. The origin of these is not known.

Several lower brominated PBDEs, including BDE-47, -99 and -100,have been shown to activate and inhibit the Ah receptor. Bromkal 70 5DE,a PeBDE product, has been shown to induce Ah-receptor mediated liverenzymes such as EROD, in vitro as well as in vivo in rats and rainbow trout.BDE-47 and -99 have been shown to decrease EROD activity in rainbowtrout liver however. Hydroxylated PBDE and TBBPA are potent competi-tors for transthyretin, the plasma protein responsible for transporting thy-roid hormones in the plasma. Brominated structural analogues of thyroxineand T3 also interact with thyroid hormone receptors. Rats and mice treatedwith Bromkal 70 5DE or BDE-47 had decreased thyroxine levels as wellas changes in immune response. Oral administration of BDE-47 or BDE-99 to neonatal mice on day 10 induced permanent aberrations in spontane-ous motor behavior which worsened with age. Neonatal exposure to BDE-99 also affected learning and memory functions in the adult animal.

Therefore, these substances have a potential to induce/down-regulateliver enzyme production, negatively influence the regulation of the thyroidhormone system and induce immunotoxicity. They also induce neurotox-icity when administered at a sensitive period of brain growth.

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Based on levels of PBDE in fatty Baltic Sea fish and current toxicolog-ical studies, the Swedish Food Administration has performed a risk assess-ment for human consumption. Based on this, they reached a no effect level(NOEL) of 2 mg/kg body weight/dag (Darnerud et al., 1998).

The time trends studied all indicate increased levels of PBDEs in the en-vironment since the 1970s. The trends in guillemot indicate that levels ofBDEs -47, -99 and -100 have begun to decline in the Baltic Sea since vol-untary withdrawal of use in a number of countries, but that levels of HBCDare increasing. The temporal trend in pike from Lake Bolmen, indicates alevelling off of PBDE levels. However, the human breast milk trend indi-cates that levels are increasing exponentially, doubling every five years.These differences may reflect differences in exposure routes. Lake Bolmenreceives its input of contaminants from long-range transport in air. TheBaltic Sea is affected by both long-range transport as well as direct releasesof contaminants into the environment from cities and agriculture. These re-sults may indicate that humans are exposed to these substances not justfrom the diet, which is the common denominator with pike and guillemot,but also from current exposure to electronic appliances and textiles con-taining brominated flame retardants in the home and work environment.

PBDE, TBBPA and HBCD are present in the environment. They are tak-en up by living organisms, and lower brominated PBDE biomagnify. TBB-PA and PBDE and/or their metabolites have been shown to be biologicallyactive. Levels of PBDE seem to be increasing, and the trend in humans inparticular indicates that this increase may be rapid. This could lead to lev-els high enough in wildlife and/or humans to be causing effects. However,knowledge about these substances, their sources, toxicity and environmen-tal behavior is very limited, making risk assessment and remedial actiondifficult. The data base for TBBPA and HBCD is so limited that it is almostimpossible to estimate risk. These results indicate that brominated flameretardants may be a new ”PCB problem”.

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RAPPORT 4697

REPORT 5065

REPORT 5065

Brom

inated Flame R

etardantsR

EP

OR

T 5

06

5

POP

POP

POPpersistent organic pollutants

isbn 91-620-5065-6issn 0282-7298

swedish environmental protection agency

Brominated

Brominated

Brominated Flame RetardantsFlame Retardants

Flame Retardants

Flame Retardants

The fact that several brominated flame retardants (BFRs)have been found in the environment, sometimes at increasinglevels, has caused great concern in many fora.

The time trends studied, indicate increased levels of manyBFRs in the environment since the 1970s but the levels ofimportant lower-brominated compounds have begun todecline in the Baltic Sea since voluntary withdrawal of use ina number of countries. However, the human breast milktrend indicates that levels are increasing exponentially,doubling every five years.

The results may indicate that humans are exposed to thesesubstances not just from the diet, but also from exposure toelectronic appliances and textiles.

These findings, and many more, are discussed in this report.The results indicate that BFRs may be a new "PCB problem".

Brominated

Brominated

Brominated

Flame Retardants

Flame Retardants

Flame Retardants

Cynthia A. de Wit

RAPPORT 5065 00-07-03, 10.331