Flame retardant concentrations and profiles in wild birds associated

with landfill: A critical review

Andrew D.W. Tongue a,b, S. James Reynolds a,c, Kim J. Fernie b,d, Stuart Harrad b, *

a Centre for Ornithology, School of Biosciences, College of Life & Environmental Sciences,

University of Birmingham, Edgbaston, Birmingham B15 2TT, UK

b School of Geography, Earth and Environmental Sciences, College of Life & Environmental

Sciences, University of Birmingham, Edgbaston, Birmingham B15 2TT, UK

c The Army Ornithological Society (AOS), c/o Prince Consort Library, Knollys Road, Aldershot,

Hampshire GU11 1PS, UK

d Ecotoxicology and Wildlife Health Division, Wildlife and Landscape Science Directorate,

Environment & Climate Change Canada (ECCC), Burlington, ON L7S 1A1, Canada

*

Corresponding author, School of Geography, Earth and Environmental Sciences, College of Life

& Environmental Sciences, University of Birmingham, Edgbaston, Birmingham B15 2TT, UK.

E-mail address: s.j.harrad@bham.ac.uk (S. Harrad).

Keywords:

Ecotoxicology; Legacy Brominated Flame Retardants; Novel Brominated Flame Retardants

(NBFRs); Organophosphate Flame Retardants (PFRs); Dechlorane Plus (DDC-CO)

Declarations of Interest: None.

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ABSTRACT

Given factors such as their persistence and toxicity, legacy brominated flame retardants (BFRs)

like polybrominated diphenyl ethers (PBDEs) and hexabromocyclododecane (HBCDD), are

designated as persistent organic pollutants (POPs) and are subject to regulation. Waste streams

likely represent a substantial reservoir of legacy BFRs given that they were once widely applied

to goods which are increasingly likely to be obsolete. Waste streams are also increasingly likely

to be a source of emerging flame retardants, in particular, novel BFRs (NBFRs), the halogenated

norbornene flame retardant Dechlorane Plus (DDC-CO) and the brominated, chlorinated or non-

halogenated organophosphate triester flame retardants (PFRs). Many bird populations rely on

landfill and its surrounding land-use for inter alia the opportunities it provides for activities such

as foraging and resting. However, studies on captive and wild (free-living) birds have

demonstrated deleterious effects of several FRs. Globally, approximately 250 bird species,

including many of conservation concern, are reported to use landfill and surrounding habitat

(including wastewater treatment operations), thus putting birds potentially at risk of exposure to

such chemicals. We synthesise and critically evaluate a total of 18 studies covering eight avian

species published between 2008 and 2018 (inclusive) across four continents that report flame

retardant (FR) burdens in birds utilising landfill. Several such studies found FRs at among the

highest concentrations detected in wild biota to date. We recommend that ongoing research be

focused on landfill-associated birds, given that landfill is an important source of FRs and other

anthropogenic chemicals, and particularly at sites where species are of conservation concern. We

suggest ways in which the comparative power of studies could be enhanced in the future, the

reporting of a minimum common suite of key chemicals, and where feasible, standardisation of

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the tissue compartments (i.e., eggs) to be studied. We conclude by identifying future research

directions.

1. Introduction

Brominated flame retardants (BFRs) have been used extensively worldwide in recent

decades, largely as additives to comply with fire safety standards and regulations. Manufacture

and use of BFRs grew in response to rapidly increasing global levels of polymeric domestic and

industrial products from the 1970s onwards (Alaee et al., 2003; Besis and Samara, 2012;

Darnerud, 2003). BFRs were incorporated into many products, including circuit boards, wire and

cable insulation, electronic and electrical goods, as well as textiles, soft furnishings, building

materials and paper and wood-based products (de Wit et al., 2002; Jenssen et al., 2007;

Stubbings and Harrad, 2014; Wu et al., 2008). Over 70 commercially-used BFRs are in

circulation. The most widely employed in recent decades are detailed in Table 1 and comprise

tetrabromobisphenol-A (TBBPA), hexabromocyclododecane (HBCDD) and three technical

mixtures of polybrominated diphenyl ethers (PBDEs): penta-BDE, octa-BDE and deca-BDE

(Alaee et al., 2003; BSEF, 2003; Leonards et al., 2007; Stubbings and Harrad, 2014). Despite

their attractive properties to industry, environmental concerns over PBDEs emerged in the late

1990s, followed by HBCDD in the mid-2000s. These concerns centered around the resistance of

these chemicals to degradation, their toxicity, capacity for long-range atmospheric transport and

bioaccumulative properties (Chen and Hale, 2010; Darnerud, 2003; de Wit et al., 2002; Law et

al., 2003). Their adverse effects on birds have been studied through investigations of various

endpoints in free-living and captive avian subjects (Guigueno and Fernie, 2017).

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The lower brominated tri- to octa-BDEs were designated as persistent organic pollutants

(POPs) by the United Nations Environment Programme (UNEP) under the Stockholm

Convention in 2009 (UNEP, 2009; Chen and Hale, 2010; Stubbings and Harrad, 2014; ECHA,

2015). They have been banned in the European Union (EU) since 2004, at which time in the

United States (US), producers agreed to a voluntary phase-out (La Guardia et al., 2006; Renner,

2004). HBCDD was listed as a POP in November 2014 (Stockholm Convention on Persistent

Organic Pollutants, 2017), with manufacturers discontinuing its production shortly thereafter

(Edser, 2015). The fully brominated deca-BDE (composed predominantly of the BDE congener

209) undergoes debromination and degradation during waste treatment and disposal, and as a

result of metabolism in organisms; it has also been banned in the EU since 2008 (European Court

of Justice, 2008). The two US producers of deca-BDE ceased production, importation and sale

for all uses domestically in 2013 (United States Environmental Protection Agency, 2010). Deca-

BDE was listed under the Stockholm Convention in May 2017. Some legacy BFRs have not

been completely phased out globally. For example, Shi et al. (2018) report that deca-BDE and

HBCDD are still produced in China.

Various alternative FRs have entered widespread use in response to these developments,

including several ‘novel’ brominated flame retardants (NBFRs), organophosphate triester FRs

(PFRs) and Dechlorane Plus (C18H12Cl12; DDC-CO). Harju et al. (2009) estimated that

approximately 100,000 metric tonnes per year of 21 NBFRs are produced globally. In terms of

PFR usage and production, van der Veen and de Boer (2012) provided global production figures

in 1997 for tris (2-chloroisopropyl phosphate) (TCIPP) of 22,950 metric tonnes (hereafter

‘tonnes’) and for (tris(2-chloroethyl) phosphate) (TCEP) of 2,040 tonnes. They also reported

production figures of certain other PFRs for some countries. Olukunle et al. (2018) report that

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the chlorinated FR, Dechlorane Plus (DDC-CO) is currently produced at two sites globally,

namely Niagara Falls, NY, US and Huai’an Province, China. In 2011, it was listed as a high-

production volume chemical in the US (i.e., it is produced or imported in quantities of at least

450,000 kg per year) (Sverko et al., 2011). However, we have been unable to find data on the

global production of DDC-CO.

Most FRs are applied to products via one of two methods: ‘additive’, where the commercial

FR mixture is blended with the molten polymer, and ‘reactive’ where FRs are covalently-bound.

PBDEs and HBCDD fall into the former category, with TBPPA generally (but not exclusively)

used as a reactive FR (Harrad and Stubbings, 2014). Additive FRs are known to leach into the

environment from the products containing them (Hutzinger and Thoma, 1987). However, despite

recent advances in our understanding of their environmental fate and toxicity, several authors

have identified knowledge gaps in terms of the impacts of emerging FRs on biota (Covaci et al.,

2011; Ezechiáš et al., 2014; Greaves and Letcher, 2017; Sverko et al., 2011; van der Veen and de

Boer, 2012).

Birds are among the more vulnerable animal classes in terms of the general toxicity of POPs

(Chen and Hale, 2010; Fernie et al., 2017; Guigueno and Fernie, 2017). Compared to mammals,

for example, birds possess generally less effective detoxification mechanisms for many

pollutants, including organohalogens (Newton, 1986, 1998; Marteinson et al., 2011). The

impacts of organohalogens on birds became widely publicised in Europe and North America in

the mid-20th Century when it was demonstrated that chlorinated hydrocarbon pesticides, most

notably dichlorodiphenyltrichloroethane (DDT) and its metabolite

dichlorodiphenyldichloroethylene (DDE), had multiple negative effects on raptors, especially

that of eggshell thinning which significantly reduced their breeding performance through

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dramatic reductions in hatching success. Raptors were also shown to be at risk of poisoning via

the bioaccumulation of cyclodiene pesticides (i.e., aldrin, dieldrin and heptachlor epoxide) in

their prey (Newton and Bogan, 1974; Ratcliffe, 1967). Being lipophilic, organohalogens

sequester in the yolk compartment of eggs and therefore the analysis of egg contents is a

relatively convenient and non-invasive method of exposure assessment and environmental

monitoring (Chen and Hale, 2010; Furness and Greenwood, 1993; Guigueno and Fernie, 2017).

Birds are important bioindicators of wider environmental pollution (Furness and Greenwood,

1993; LeBlanc and Bain 1997; O’Brien et al., 1993). Several long-running egg-sampling

(‘specimen-banking’) studies exist globally. Examples include the North American Great Lakes

Herring Gull Monitoring Program (Hebert, 1999), the UK’s Predatory Bird Monitoring Scheme

(Walker et al., 2008) and the German Environmental Specimen Bank (Paulus et al., 2010;

Koschorreck et al., 2015). More recently (i.e., between 2010 and 2015), European raptor

sampling was coordinated via the EURAPMON initiative (Gómez-Ramírez et al., 2014).

In this review, we define ‘landfill’ as sites designated for the practice of domestic and

commercial waste disposal, primarily at designated sites serving a municipality (i.e., municipal

solid waste [MSW]), but also including the informal dumping of any waste incorporating

anthropogenic food refuse. Landfill appears to constitute an important reservoir of environmental

legacy BFRs, particularly as it is a repository of obsolete goods treated with these formerly

widely used chemicals (Danon-Schaffer and Mahecha-Botero, 2010; Stubbings and Harrad,

2014). The presence of NBFRs in waste streams has been reported previously (Covaci et al.,

2011; McGrath et al., 2017; Monclús et al., 2018; Li et al., 2017; Olukunle and Okonkwo, 2015;

Nyholm et al., 2013). Landfill is also an important resource for various wild vertebrates,

including birds (Oro et al., 2013; Plaza and Lambertucci, 2017). To date, limited attention has

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been paid by toxicologists, ornithologists and wildlife conservationists to landfill as a potentially

important source of organohalogen contamination in birds. In this review we demonstrate that

birds associated with landfill can exhibit elevated FR burdens compared to those birds that do

not utilise landfill (e.g., Chen et al., 2013; Gentes et al., 2015; Muñoz-Arnanz et al., 2011a,

2011b, 2012).

2. Landfill as a source of environmental FR contamination

Among the routes by which anthropogenic chemicals enter the environment, waste disposal

ranks amongst the most important (Danon-Schaffer et al., 2013; De Lapuente et al., 2014; Eggen

et al., 2010; Masoner et al., 2015; Teuten et al., 2009; Weber et al., 2011). Landfill has been

identified as a source of abiotic environmental FR contamination on various continents,

including Africa (Odusanya et al., 2009; Daso et al., 2017), Asia (Eguchi et al., 2013; Hafeez et

al., 2016; Ilyas et al., 2010; Kwan et al., 2013; Osako et al., 2004), Australasia (Gallen et al.,

2016), Europe (Morris et al., 2004; Morin et al, 2017) and North America (Danon-Schaffer et al.,

2006; Gavilan-Garcia et al., 2017; Oliaei et al., 2002). McGrath et al. (2017) compared landfill

versus reference soils in five Asian countries and reported a significantly higher level of PBDE

soil contamination at landfill sites. In Vietnam, mean concentrations of ∑14 PBDEs in landfill

were 95 ng/g compared to 0.22 ng/g at a reference site (Eguchi et al., 2015). The pathways by

which BFRs transfer from waste products to the environment were reviewed by Stubbings and

Harrad (2014). Two main pathways were identified: emissions to air and leaching to

groundwater, with both pathways facilitated by abrasion of polymeric waste (releasing particles

to air and leachate), and photolytic and/or biodegradation of less volatile and water-soluble high

molecular weight FRs (such as deca-BDE) to lower brominated products.

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Landfill is the most common method of waste disposal worldwide, despite reforms aimed at

eliminating it in some jurisdictions such as the EU (European Union, 1999). Landfill receives

annually approximately 59% of total global MSW (World Bank, 2012). Global solid waste

generation is anticipated to peak later this century (European Union, 1999, Hoornweg et al.,

2013; Eggen et al., 2010). Alcock et al. (2003) estimated that over 80% of total BFR-containing

waste in the UK and North America enters landfill, with the remainder incinerated. Danon-

Schaffer et al. (2013) predicted that even an immediate ban on BFR-treated products entering

landfill would not result in a decline in landfill PBDE concentrations until 2080.

3. Bird utilisation of landfill

The United States Environmental Protection Agency (US EPA) estimated that domestic food

waste in 2010 comprised 13.9% (31.5 million tonnes) of total MSW (United States

Environmental Protection Agency, 2011). Parfitt et al. (2010) suggested that 30-40% of global

food designated for human consumption is wasted, with the UK, for example, disposing of 14

million tons of food and drink annually. Such waste is regularly available in space and time,

thereby constituting a predictable anthropogenic food source for wildlife, including birds (Oro et

al., 2013; Osterback et al., 2015, Plaza and Lambertucci, 2017; Oka et al., 2016). The enclosed

fringes of some landfill facilities may also contain relatively undisturbed terrestrial and aquatic

habitats that are used as foraging and breeding sites for various avian taxa (Mehra et al., 2017;

Tuljapurkar and Bhagwat, 2007).

A search of the online databases Google Scholar, Medline, Scopus and Web of Science using

‘birds AND landfill OR dump OR rubbish OR refuse OR garbage’ as search terms with Boolean

operators identified a total of 247 avian species recorded globally at municipal solid waste

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landfill sites. Two of these studies also incorporated sewage treatment facilities (Table S1 in

Supplementary Data). This total comprises a diversity of species, including those that forage

directly on anthropogenic waste (e.g., gulls [Laridae]; Fig. 1), those that utilise aquatic and

vegetated habitats within landfill sites (e.g., wildfowl, shorebirds, songbirds), plus those that may

target landfill as a source of live prey (e.g., raptors). Eighteen of the identified species are

designated by the International Union for Conservation of Nature (IUCN) as being of

conservation concern. Aside from the foraging opportunities that it affords birds, landfill sites

may be important for some groups such as gulls for loafing (i.e., resting or preening). Herring

gulls (Larus argentatus) have been found to spend considerable proportions of their daily time

budgets (i.e., >60%) loafing at landfill sites (Belant et al., 1993; Patton, 1998). While some

authorities have recently divided this species into the European herring gull versus the American

herring gull, in this paper, we use the general term ‘herring gull’. Landfill can also be important

for gull social interactions, including courtship and copulation (Belant et al., 1993).

The consumption of anthropogenic waste by birds is associated with both costs and benefits

for individuals and populations. For example, consuming such items may correlate with elevated

body condition (Steigerwald et al., 2015), breeding performance (Djerdali et al., 2016; Pons,

1992; Weiser and Powell, 2010), and population size (Bosch et al., 1994; Smith and Carlile,

1993; Mazumdar et al., 2018; Rock, 2005). The year-round presence of landfill-provided food

has been linked with reduced migratory demands placed upon some species, such as the white

stork (Ciconia ciconia) (Arizaga et al., 2018; Gilbert et al., 2016) and the lesser black-backed

gull (Larus fuscus) (Barnes, 1961) in Europe. Species of conservation concern may benefit from

the regular availability of such food (Jaksic et al., 2001). However, other studies report negative

impacts of foraging at landfill, such as reduced breeding performance (Belant et al., 1998;

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Katzenberger et al., 2017; O’Hanlon et al., 2017; Pierotti and Annett, 1991; Real et al., 2017)

and impaired fitness (Annet and Pierotti, 1999). Stable isotope analysis of feathers taken from

museum skins by Hobson et al. (2015) suggested that glaucous-winged gulls (Larus glaucescens)

at the Salish Sea (Canada and US) are increasingly reliant on human refuse and terrestrial

invertebrates, instead of marine foods. This has correlated with declines in egg volume and

clutch size in this species since the early 20th Century (Blight et al., 2015). In addition, birds

associating in large aggregations at anthropogenic waste facilities can increase the prevalence

and spread of pathogens in, and among, individuals (Macdonald and Standring, 1978; Ortiz and

Smith, 1994; Elliott et al., 2006; Nicastro et al., 2018).

4. FR contamination in bird populations associated with landfill

We identified 18 peer-reviewed studies that reported FR concentrations in bird populations

known, or suspected, to forage on landfill (Table 2). These are reviewed below in order of

publication date, except for research on gulls and starlings, which are grouped separately

(ordered by publication date). To aid comparison between studies, we express concentrations in

wet weight (ng/g ww), unless otherwise stated. Table 3 compares PBDE burdens in landfill and

reference birds of the same species. In addition to the studies identified, it should be noted that a

broad suite of FRs have been found in the eggs of herring gulls in the North American Great

Lakes Basin (Hebert, 1999) and in Germany (Koschorreck et al., 2015), as well as in Arctic-

breeding glaucous gulls (Larus hyperboreus) (Verreault et al., 2018). These authors have

suggested that use of landfill for foraging is one likely source of such contaminants in these

birds.

The first study to make an explicit connection between elevated FR burdens of birds and

their association with landfill was Polder et al. (2008). In this study, PBDE and HBCDD

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concentrations were reported in the eggs of eight species, albeit with several low sample sizes (n

= 1–20). Sum (Σ)8PBDEs and total (i.e., α, β and γ) HBCDD concentrations were highest in

African sacred ibis (Threskiornis aethiopicus) eggs (n = 2) (ranges of 4.03–19.8 ng/g and 0.31–

3.5 ng/g ww, respectively). The authors suggested that elevated BFR concentrations in this

species likely resulted from the birds supplementing their natural diet of fish, amphibians and

invertebrates with anthropogenic waste obtained from landfill sites. Future studies of legacy BFR

contamination in birds in South Africa associated with landfill would benefit from increased

sample sizes where possible.

Muñoz-Arnanz et al. (2011a) collected addled (failed) eggs of white storks to investigate

PBDE profiles and burdens in two contrasting Spanish colonies: one located in urban-industrial

Madrid (n = 10) where birds fed predominantly on landfill, the other in the predominantly rural

Coto-Doñana National Park (n = 23) where birds had a more natural diet consisting of crayfish

(Decapoda spp.), terrestrial invertebrates, earthworms (Lumbricidae) and amphibians. Urban

eggs had approximately six times higher mean ∑28PBDE concentrations (mean: 9.08, range:

2.79–20.5 ng/g ww) than rural eggs (mean: 1.64, range: 0.214–9.50 ng/g ww). The relative

contribution of BDE-209 to ∑PBDE concentrations was significant for both urban (44.1%) and

rural (38.6%) sites. Muñoz-Arnanz et al. (2011a) measured BDE-202 concentrations in all urban

eggs and in approximately 80% of rural eggs. Since this congener has not been reported in any

commercial PBDE mixture, its detection was likely either due to debromination in the

environment and subsequent uptake, or as a metabolic product of higher brominated BDE

congeners (i.e., BDE-209), as demonstrated in captive American kestrels (Falco sparverius)

exposed to BDE-209 (Letcher et al., 2014). The same stork eggs were analysed for DDC-CO

concentrations (Muñoz-Arnanz et al., 2011b) which revealed that urban eggs contained

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significantly greater total DDC-CO median concentrations (0.306 ng/g ww) compared with rural

eggs (0.085 ng/g ww).

It is well documented that ring-billed gulls (Larus delawarensis) in North America utilise

urban-industrial environments, including landfill and wastewater treatment plants. Use of such

facilities by this species near Montreal (Canada) may explain the elevated FR concentrations in

birds breeding at a major colony near that city (Ille Deslauriers). Patenaude-Monette et al. (2014)

demonstrated the importance of waste as a food source for this colony, with over 40% of the

birds’ diet comprised of anthropogenic food refuse. Gentes et al. (2012) reported mean

∑45PBDEs in livers of these birds of 205 ± 32 ng/g ww; range: 22.4–682 ng/g ww. In a pan-

Canadian study of FRs in the eggs of gulls breeding at 26 colonies, Chen et al. (2012a) found

that the highest BFR burdens were in herring gulls and ring-billed gulls breeding at Ille

Deslauriers (median ∑16PBDE concentrations of 610 and 144 ng/g ww, respectively). Using

GPS telemetry, Gentes et al. (2015) found that ring billed gulls that visited waste management

facilities also had a smaller relative plasma percentage of the lower brominated penta-BDE

congeners (i.e., the sum of BDEs -47, -99 and -100) associated with a higher trophic level diet

(e.g., Roscales et al., 2016). Ille Deslauriers-breeding ring-billed gulls exhibited mean total

HBCDD liver concentrations of 5.22 ± 1.02 ng/g ww; n = 28) (Gentes et al., 2012). Gentes et al.

(2015) also demonstrated that only brief visits (<5 % of time away from this colony, lasting

approximately 40 minutes) by ring-billed gulls to landfill and wastewater treatment plants

resulted in the uptake of significant concentrations of deca-BDE. Conversely, plasma deca-BDE

concentrations were not related to the percentage of time gulls spent in agricultural fields, urban

areas and on the St Lawrence River. Male ring-billed gulls that visited waste management sites

contained a greater percentage of BDE-209 in plasma relative to those that did not (38 ± 5 %

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versus 19 ± 1 %[SE]) (Gentes et al., 2015). Among the NBFRs detected in the livers of Ille

Deslauriers-breeding ring-billed gulls, Gentes et al. (2012) reported what is believed to be the

highest detection frequency (89% of samples) and highest concentration (max: 17.6 ng/g ww) of

the NBFR, bis(2-ethylhexyl)-2,3,4,5-tetrabromophthalate (BEHTBP), found in free-living birds

to date. Ring-billed gulls from Ille Deslauriers also exhibited one of the highest maximum total

(syn- and anti-) DDC-CO egg concentrations found across Canada (2.4 ng/g ww) (Chen et al.,

2012a). Collectively, the findings of these studies raise questions about the exposure to FRs of

migratory species of conservation concern that may use waste management facilities as

temporary stop-over sites (Mehra et al., 2017; Moulaï, 2007; Tuljapurkar and Bhagwat, 2007).

It is well known that organohalogen burdens may be lower in female birds compared to

males, likely due to a proportion of contaminant loads being deposited in eggs. In landfill-

foraging birds such differences may be also influenced by competition between males and

females. For example, in direct competition for food items obtained by terrestrial foraging on

landfill, male herring gulls competitively exclude females (Monaghan, 1980; Greig et al., 1984).

This may result in males having higher FR burdens. Gentes et al. (2015) found that male ring-

billed gulls exhibited greater ∑PBDE plasma concentrations (38.5 ± 5.0 ng/g ww) than females

(24.4 ± 3.2 ng/g ww), as well as greater relative deca-BDE concentrations, despite GPS

telemetry showing that both sexes spend similar amounts of time on landfill (Gentes et al., 2015).

In terms of PFRs in bird species known to utilise landfill, Greaves et al. (2016a) reported

concentrations of TCEP, TCIPP, triphenyl phosphate (TPHP) and tris(2-butoxyethyl) phosphate

(TBOEP) ranging from ND to 6.69 ng/g ww in a retrospective analysis of pooled egg

homogenates from five Great Lakes herring gull colonies collected between 1990 and 2010.

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Landfill was postulated as an origin of these PFRs, although natural food (fish) from the aquatic

food web could not be excluded as a source.

Several Mediterranean-based studies (e.g., Bertolero et al., 2015; Morales et al., 2012, 2016;

Pastor et al., 1995; Roscales et al., 2016; Zapata et al., 2018) have compared the burdens and

profiles of various contaminants, including FRs, in the eggs of the marine piscivore, Audouin’s

gull (Ichthyaetus audouinii), with those of the generalist (and landfill-associated) yellow-legged

gull (Larus michahellis). The two species regularly occur in mixed colonies (González-Solís et

al., 1997). Morales et al. (2012) found a predominance of BDE-209 in both species breeding in

the industrial Ebro Delta (north-eastern Spain). However, in the south-western Mediterranean

Sea, Roscales et al. (2016) reported that the higher trophic level diet of the Audouin’s gulls was

the likely source of a higher percentage of lower brominated BDE congeners (i.e., tetra- and

penta- BDEs) than in the yellow-legged gulls. Conversely, the latter species laid eggs containing

higher percentages of higher brominated congeners (i.e., hepta-, octa- and deca- BDEs).

Sardines (Sardina pilchardus) are an important prey species for Audouin’s gulls (Witt et al.,

1981). Parera et al. (2013) reported levels of ∑PBDEs of 5.49 ng/g lipid weight (lw) in sardines,

with lower brominated congeners (BDEs -28, -47 and -100) predominating. Fishes appear to be

an important dietary source of more bioavailable lower brominated PBDE congeners in both

marine and freshwater piscivorous avifauna. Similarly, elevated BDE burdens (>1000 ng/g ww)

were reported in the eggs of ospreys (Pandion haliaeetus), an obligate piscivorous raptor, in

Oregon and Washington State, US (Henny et al., 2009). As in the case of Audouin’s gulls, lower

brominated BDE congeners predominated in these osprey eggs: BDE-47 was the dominant

congener followed by BDEs -100, -99, -154 / BB153, -153 and -28 (Henny et al., 2009).

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Birds with a terrestrially-based diet appear to accumulate greater levels of DDC-CO than

those with predominantly aquatic diets (for example, yellow-legged gulls versus Audouin’s gulls

in Spain; herring gulls versus Caspian terns in the US) (Muñoz-Arnanz et al., 2012; Su et al.,

2017, respectively). Given the association of both yellow-legged gulls and herring gulls with

landfill, landfill facilities may be a specific source of DDC-CO in such birds. As mentioned

above, Madrid-breeding white storks, known for their use of landfill, exhibit higher DDC-CO

concentrations than conspecifics breeding in the rural Coto-Doñana.

Common starlings (Sturnus vulgaris) breeding at Canadian landfill sites showed orders of

magnitude higher ∑PBDE egg concentrations (range: 11–805 ng/g ww) compared to those

breeding at adjacent urban centres (range: 4.9–146 ng/g ww), and agricultural locations 10 km

and 40 km distant (ranges: 1.9–488 and 2–375 ng/g ww, respectively) (Chen et al., 2013).

Starlings routinely forage at landfill sites (Belant et al., 1995; Burger and Gochfeld, 1983). Chen

et al. (2013) reported maximum egg PBDE concentrations laid by landfill-associated starlings

that were among the highest ever-reported in free-living passerines and comparable to egg

concentrations reported in some terrestrial raptors (e.g., 616 ng/g lw in liver of Eurasian

sparrowhawk (Accipiter nisus) in Belgium; Voorspoels et al., 2007). Chen et al. (2013) reported

that the median Σ14PBDE egg concentrations of starlings breeding on landfill and the human

population density of the surrounding area were significantly and positively correlated (r = 0.99,

p < 0.001 n = 7). BEHTBP was also found in the eggs of landfill-breeding starlings (maximum

concentration: 26 ng/g ww). Total (syn- and anti-) DDC-CO was reported in lower

concentrations than PBDEs in these starling eggs across all land-use categories.

In contrast to most other studies described in this review, Chen et al. (2013) found that BDE-

209 accounted for a small proportion (1–15%) of ΣPBDE concentrations in starling eggs, with no

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relationships observed with land use. It should be noted that Van den Steen et al. (2007)

concluded that BDE-209 in blood and tissues of captive common starlings was either eliminated

or debrominated to lower brominated BDE congeners. In terms of other FRs, Lu et al. (2017) did

not find elevated PFR concentrations in the eggs of starlings in proximity to landfill across the

whole of Canada, despite establishing that (non-FR) volatile methylsiloxanes (VMS) in landfill-

associated starling eggs contained significantly higher Σ6VMS concentrations (median: 178 ng/g

ww) (n = 29) than those laid by birds at urban-industrial (20 ng/g ww) (n = 16) and rural

locations (1.3 ng/g ww) (n = 26). However, PFRs have been shown to metabolise rapidly in

herring gulls (Greaves and Letcher, 2014), bald eagles (Haliaeetus leucocephalus) (Stubbings et

al., 2018) and American kestrels (Fernie et al., 2015).

Other studies report elevated BFR concentrations in free-living common starling populations

associated with landfill. Comparing PBDE (and other contaminant) egg burdens of common and

spotless starlings (Sturnus unicolor) in 13 countries across three continents, Eens et al. (2013)

found common starling eggs (n = 10) at a municipal landfill in Vancouver, BC, Canada to

contain the highest ∑7PBDE concentrations (4400 ± 830 ng/g lw), compared to the lowest of 3.7

± 1.1 ng/g lw [SE] in spotless starling eggs (n = 10) at a rural location in Spain. BDE-99

accounted for 55% of ∑PBDEs in eggs from the Vancouver landfill, likely due to its presence in

the penta-BDE commercial formulation, which has been used more extensively in North

America than elsewhere (BSEF, 2003; Table 1). At the same Vancouver landfill, starling

nestlings contained ∑6PBDE plasma concentrations approximately 60 times in excess of plasma

from nestlings at a rural reference site 40 km from the city (Erratico et al., 2015).

Black kites (Milvus migrans) forage on landfill in various countries, including India

(Mazumdar et al., 2016), Italy (Da Giacomo, 2008), Spain (Blanco, 1997, 2018), Turkey (Biricik

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and Karakaş, 2011) and Uganda (Pomeroy, 1975). Their generalist diet and close association

with urban centres was suggested to be the likely reasons for their elevated FR concentrations (n

= 8) compared with those of nine other avian taxa in Pakistan (Abbasi et al., 2016a). Black kites

had median tail feather concentrations of ∑7PBDEs of 2.4 (range: 0.70–7.50) ng/g dry weight

(dw) and of total HBCDD concentrations of 1.5 (range: 0.5–8.1) ng/g dw. Abbasi et al. (2016a)

found that this was the only species in which concentrations of 1,2-bis(2,4,6-tribromophenoxy)

ethane (BTBPE) were quantifiable, albeit at low levels (median: 0.10 (range: <LOQ-0.1) ng/g

dw; BTBPE LOQ 0.02 ng/g dw). Blanco et al. (2018) measured concentrations of PBDEs and

dechloranes in addled eggs of black kites breeding in urban Madrid (n = 28) and at Coto-Doñana

(n = 23). Madrid-breeding black kites are known to forage on landfill, whereas those breeding at

the Coto-Doñana have a more natural diet which includes waterbirds (Blanco et al., 2018). Black

kite eggs from Madrid showed significantly higher PBDE concentrations than those from

Doñana (maximum ∑PBDEs: 1570 ng/g lw versus 13.6 ng/g lw, respectively). However, eggs

from Doñana contained higher concentrations of DDC-CO (9.34 ± 7.56 [± 1 SD] versus 0.55 ±

1.46 ng/g lw) which might be explained by the greater bioaccumulation of DDC-CO in the

natural prey of Doñana black kites and / or possible sources of DDC-CO relating to industrial

activity or waste management facilities. Monclús et al. (2018) observed that adult cinereous

vultures (Aegypius monachus) also foraged at a landfill close to Madrid. Median ∑4PFRs (TCEP,

TCIPP, TPHP and tris (1,3-dichloro-2-propyl) phosphate [TCIPPP]) concentrations in down

feathers of vulture nestlings (20.6 ng/g dw) were several orders of magnitude higher than median

∑7PBDEs (0.51 ng/g dw) (n = 16). This pattern suggests that the landfill in question could be a

richer source of PFRs than of legacy BFRs, a finding which may potentially be more frequently

encountered as time progresses following BFR phase-out.

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5. Routes of exposure, toxicokinetics and toxicity of FRs in landfill-associated birds

Birds may be exposed to FRs and other POPs through several routes, including diet, air and

water, although more research is required to better characterise these. Landfill-foraging gulls, for

example, are known to ingest not only putrescible organic matter, but also plastics, metals, glass,

building material, rubber and fabrics (Basto et al., 2019; Bond, 2016; Seif et al., 2017). This is

likely to result from their highly competitive foraging behaviours in large aggregations, where

competition is intense and foraging opportunities are short-lived as a result of their regular

displacement by other birds and heavy machinery (Coulson, 2015). Dermal exposure is an area

of increasing interest in the field of human BFR toxicokinetics (Abdallah et al., 2018) but it is an

under-researched area in avian toxicology (Alharbi et al., 2016; Mineau, 2011). Airborne

particulates have been shown to be an important contaminant pathway in birds (Sorais et al.,

2017; Fernie et al., 2018) and also warrant further study, ideally in combination with GPS

telemetry and other technology, to determine the spatio-temporal aspects of avian use of landfill

in relation to contaminant burdens (Gentes et al., 2015; Gyimesi et al., 2016; Navarro et al.,

2016; Rock et al., 2016).

The impacts of BFRs on birds have been reported for a range of captive and free-living

species (recently reviewed in Guigueno and Fernie, 2017). However, this remains a nascent field

and considerably less is known about the impacts of NBFRs and other emerging FRs (Ezechiáš

et al., 2014; Sverko et al., 2011; van der Veen and de Boer, 2012). The effects of FRs on

landfill-foraging birds could be amplified if birds are exposed to complex mixtures involving not

only FRs, but also other anthropogenic chemicals known to occur at waste disposal facilities,

such as polychlorinated biphenyls (PCBs), polycyclic aromatic hydrocarbons (PAHs) (Melnyk et

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al., 2015), dioxins and furans (Gómez-Lavín et al., 2018), and trace elements (de la Casa-Resino

et al., 2014). Recent studies show that landfill-associated common starlings in Canada are also

exposed to perfluoroalkyl acids (Gewurtz et al., 2018) and volatile methylsiloxanes (Lu et al.,

2017).

5.1. Polybrominated diphenyl ethers (PBDEs)

PBDE concentrations in avian taxa appear to be influenced by several factors. These include

diet and trophic level, exposure to different commercial mixtures and distance to point sources,

metabolic capacity, nutritional state, age, sex, and migratory behaviour (Chen and Hale, 2010;

Kelly et al., 2007; Polder et al., 2008; Roscales et al., 2016). Given that the bioaccumulation and

biomagnification potential of different BDE congeners is influenced by their octanol-water

partition coefficient (KOW), the nature of the food web concerned (i.e., aquatic versus terrestrial)

is important. Congeners with log KOW values ranging between ~5.9 and ~7.2 (i.e., tetra- and

penta- BDEs) have the greatest biomagnification potential. Biomagnification potential declines

significantly for chemicals such as octa- through deca- BDEs with log KOW values >7.2 (reduced

absorption rates) or <5.9 (efficiently eliminated by respiration) (Chen and Hale, 2010; Kelly et

al., 2007). Those avian predators most closely associated with terrestrial ecosystems (i.e., those

that predominantly consume passerines, terrestrial mammals and terrestrial invertebrates), tend to

exhibit a greater percentage of higher brominated congeners, such as BDE-209 (Voorspoels et

al., 2006). Conversely, PBDE profiles of birds at the apices of aquatic food chains are often

characterised by a greater relative percentage of more bioavailable lower brominated congeners

(Chen and Hale, 2010; Elliott et al., 2005; Law et al., 2003; Henny et al., 2009, 2011). However,

differences in such congener profiles may also reflect local use of particular commercial

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mixtures (Table 1; BSEF, 2003). For example, from the early 1980s, North America witnessed a

continuous release of penta-BDE for approximately 20 years that was associated with a relatively

rapid increase of lower brominated, more bioavailable BDE congeners in birds and other biota

(Alcock et al., 2003; Chen and Hale, 2010). Europe underwent earlier reductions in penta-BDE

use, but a lengthy period of octa- and deca-BDE utilisation may account for the comparatively

higher percentages of congeners associated with these commercial formulations reported in

European wildlife (Chen and Hale, 2010).

BDE-209 is regularly reported as the dominant PBDE congener in landfill solid matrices

(i.e., soil, sediment and biosolids) (e.g., Eguchi et al., 2013; Ilyas et al., 2010; Kajiwara et al.,

2014; Li et al., 2012; Morin et al., 2017), and this is the case across terrestrial substrates in

general (McGrath et al., 2017; Tokarz et al., 2008; Wei et al., 2016). Gavilan-Garcia et al. (2017)

did not detect BDE-209 in landfill sludge (Mexico City, Mexico) and suggested that this may

have been due to products containing this congener being yet to enter the waste stream.

However, among other congeners, BDE-209 is known to undergo sequential debromination and

degradation in abiotic matrices (Danon-Schaffer and Mahecha-Botero, 2010; Gerecke et al.,

2006; Li et al., 2018; Robrock et al., 2008; Schenker et al., 2008; Stapleton and Dodder, 2008;

Stubbings and Harrad, 2014; Tokarz et al., 2008) and has been shown to do so in birds.

Examples include free-living ring-billed gulls in Canada (François et al., 2016), and captive

American kestrels (Letcher et al., 2014) and common starlings (Van den Steen et al., 2007).

Guigueno and Fernie (2017) used an adverse outcome pathway (AOP) approach to

synthesise information from 61 in vitro and in vivo studies of the most commonly reported

endpoints in avian FR toxicity research. They reported that the most sensitive of all endpoints

was behaviour specifically related to reproduction, with significant effects identified for all such

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studies reviewed. Most relevant to the present review were the reported impacts associated with

exposure of ring-billed gulls, glaucous gulls, herring gulls and common starlings. Table S4

summarises PBDE burdens reported in landfill-associated species. In ring-billed gulls breeding at

Ille Deslauriers (Quebec, Canada), liver type-1 deiodinase mRNA expression was inversely

related to hepatic ∑octa-BDE levels (François et al., 2016). Males from this colony had bones

that were demineralised as a result of PBDE exposure (Plourde et al., 2013). In male Norwegian

Arctic-breeding glaucous gulls, plasma progesterone was positively related to concentrations of

several POPs, including PBDEs, suggesting that POPs exposure may interfere with

steroidogenesis and affect circulating progesterone homeostasis (Verreault et al., 2006).

Similarly, in the eggs of glaucous gulls breeding in Svalbard, positive relationships existed

between ∑11PBDEs and HBCDD concentrations (among other POPs) and both testosterone and

17β-estradiol (Verboven et al., 2008). In liver and brain tissue samples of glaucous gulls in

Norway and of herring gulls in Canada, hydroxylated and non-hydroxylated PBDEs, and

selected hydroxylated PCBs (polychlorinated biphenyls), disrupted triiodothyronine (T3) and

thyroxine (T4) transport via binding interactions with the protein transthyretin (TTR) (Ucán-

Marín et al., 2009). Crump et al. (2008) found ∑13PBDEs in the brains of herring gulls (n = 5) in

the Great Lakes region at concentrations of up to 143 ng/g ww. They also found down-regulation

of TTR mRNA in embryonic neuronal cells of birds collected from eastern Canada following

treatment with both T3 and BDE-71, implying a shared mechanism of action.

The effects of PBDE exposure on captive common starlings include disruption of the thyroid

system (Eng et al., 2014), elevated egg mass and volume (Van den Steen et al., 2009) and lower

T3 levels (Van den Steen et al., 2010). Evidence for PBDE debromination has been reported in

starlings (Van den Steen et al, 2007) and in other free-living avian subjects, including Eurasian

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sparrowhawks (a predator of common starlings; Crosse et al., 2012), white storks (Muñoz-

Arnanz et al., 2011a) and common kingfishers (Alcedo atthis) (Mo et al., 2012). Erratico et al.

(2015) concluded that the oxidative metabolism of PBDEs in liver microsomes of juvenile

common starlings proceeds at a slow rate which suggests that certain PBDEs are likely to

accumulate in individuals over time.

5.2 Hexabromocyclododecane (HBCDD)

While the toxic effects of HBCDD have been identified in birds in general (e.g., Guigueno

and Fernie, 2017), few studies have examined its toxicity in those bird species associated with

landfill. In eggs laid by Norwegian glaucous gulls, concentrations of both α-HBCDD and

∑11PBDEs were found to correlate with testosterone and estradiol levels (Verboven et al., 2008).

Table S5 summarises HBCDD burdens reported in seven landfill-associated species. Of the three

commonest commercial HBCDD diastereoisomers (i.e., α-HBCDD, β-HBCDD and γ-HBCDD),

the first is most commonly reported in biota because of its longer half-life and greater persistence

in organisms (Letcher et al., 2015). Landfill-associated species are apparently no exception

(Gauthier et al., 2007; Verboven et al., 2008). Global HBCDD concentrations in avifauna have

yet to show declines and, indeed, may still be rising, although data remain scarce (Abbasi et al.,

2016b; de Wit et al., 2010; Law et al., 2014; Letcher et al., 2010; Wu et al., 2012). Su et al.

(2015) reported that α-HBCDD, BDE-209 and ∑2DDC-CO concentrations in herring gull eggs

collected in 2012 under the GLHGMP were significantly higher (p < 0.05) than in 2006 and

2008 eggs collected from the same colonies, with the mean concentration of HBCDD in eggs

collected in 2012 110% greater than in those from 2006 for the same colony.

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5.3 Other FRs

NBFRs are currently reported at considerably lower concentrations in avian tissues than are

legacy BFRs. For example, Chen et al. (2013) reported BEHTBP as the only frequently

quantifiable NBFR in common starling eggs (detectable in 47% of all egg homogenates, with

concentrations ranging from <MLOD–26 ng/g ww [LOD: 0.05 ng/g ww]). In contrast, PBDEs

were found in 96% of eggs, ranging in concentration from 2–805 ng/g ww. Such observations

may reflect the relatively short time that NBFRs have been in circulation, although they might

also suggest their limited bioavailability and / or their marked biotransformation.

Of 16 PFRs in six tissue compartments of breeding female herring gulls (n=8) in the North

American Great Lakes region, alongside separate yolk and albumen analyses of their entire

clutches (n = 16), Greaves and Letcher (2014) determined that fat (32.3 ± 9.8 ng/g ww)

contained the highest ∑PFR concentrations, followed by egg yolks (14.8 ± 2.4 ng/g ww) ≈ egg

albumen (14.8 ± 5.9 ng/g ww), muscle (10.9 ± 5.1 ng/g ww) and red blood cells (1.00 ± 0.62

ng/g ww). PFRs were not detectable in liver, plasma and brain tissues. Concentrations of

different PFRs were tissue-dependent. For example, the concentration of TBOEP in albumen was

8.09 ± 2.49 ng/g ww, while in yolk it was 1.89 ± 0.69 [SD] ng/g ww. TCIPP yolk concentration

was 6.05 ± 1.36 ng/g ww, while in albumen it was 2.31 ± 1.64 (± 1 SD) ng/g ww. The presence

of PFRs in fat and muscle (i.e., long-term storage deposits), but not in liver or plasma (i.e., short-

term transient stores) suggests rapid PFR transformation to their diester metabolites. This was

demonstrated by Greaves et al. (2016b) in hepatic microsomes of herring gulls.

Toxic in vivo effects of PFRs have been documented in birds (e.g., Fernie et al., 2015). For

gulls that use landfill, the toxic effects of these particular FRs have been documented in vitro.

Embryonic hepatocytes from an unspecified number of herring gulls in the Great Lakes region

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were sensitive to the cytotoxic effects of three administered PFRs: tricresyl phosphate (TCP),

TCIPP and tris(1,3-dichloro-2-propyl) phosphate (TDCIPP), as well as the NBFR, 1,2-dibromo-

4-(1,2-dibromoethyl)cyclohexane (DBE-DBCH or TBECH) (Porter et al., 2014). However, these

were at extremely high doses (> 7000 µg/g hepatocytes) that were orders of magnitude higher

than the concentrations recorded in the eggs of conspecifics obtained from the same geographical

area (e.g., 0.0002 µg/g ww for TDCIPP; Chen et al., 2012b).

Guigueno and Fernie (2017) identified only one study which examined the effects of DDC-

CO on birds, namely domestic chickens (Gallus gallus) (Crump et al., 2011). This study found

no cytotoxicity or effects on embryonic viability at DDC-CO concentrations of up to 500 ng/g

per egg (i.e., at concentrations 10 times greater than those reported in Great Lakes herring gull

eggs). We identified NBFR, DDC-CO and PFR concentrations that have been reported in 10

landfill-associated species (tables S2 and S3, respectively).

6. Summary and research directions

This review underlines the importance of MSW landfill as a source of FR contamination in

birds that exploit this predictable anthropogenic food subsidy. We have identified 18 studies

published between 2008 and 2018 in which elevated FR burdens were found in eight bird species

(i.e., white stork, African sacred ibis, cinereous vulture, black kite, ring-billed gull, herring gull,

yellow-legged gull and common starling) regularly found on landfill in six countries (i.e.,

Canada, Pakistan, South Africa, Spain and the US). These studies analysed samples for legacy

BFRs (i.e., PBDEs and HBCDD) and / or emerging FRs (i.e., NBFRs, DDC-CO and PFRs).

Other studies have documented temporal trends in FR burdens of two gull species (i.e., herring

gulls breeding in the North American Great Lakes and the German North Sea and Baltic coasts

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plus Arctic-breeding glaucous gulls) that frequently use landfill. Several of these studies report

FR burdens in eggs at concentrations close to the highest reported in free-living biota to date,

although liver concentrations are below the maximum ever recorded in biota (a maximum

hepatic concentration of 197,000 ng/g lw in an individual Coopers hawk (Accipiter cooperi) in

Vancouver, Canada; Elliott et al., 2015). Examples of elevated egg concentrations include 2201

ng/g ww ∑14PBDEs in herring gulls in the Great Lakes region (Su et al., 2015), 23 ng/g ww total-

HBCDDs in glaucous gulls in the Norwegian Arctic (Verboven et al., 2008), 9.79 ng/g ww

NBFRs (PBEB) in white storks in Spain (Muñoz-Arnanz et al., 2010), 54.6 ng/g ww DDC-CO in

herring gulls in the US (Su et al., 2017) and 14.21 ng/g ww ∑14PFRs in the same species in

Canada (Greaves et al., 2016). The PBDE concentrations in the first example are either equal to,

or higher than the lowest observed effects of PBDEs in captive American kestrels which,

compared to control birds, exhibited various responses, including immunomodulation (Fernie et

al., 2005a), changes in thyroid, vitamin A, glutathione homeostasis and oxidative stress (Fernie

et al., 2005b), changes in reproductive courtship behaviours (Fernie et al., 2008) and reduced

eggshell thickness and reproductive success (Fernie et al., 2009).

There is a paucity of FR research in free-living birds across South America, much of Africa

and most Asian countries. Such countries are often where human populations and development,

and therefore quantities of anthropogenic waste, are increasing most rapidly (Hoornweg et al.,

2013). In countries where the funds exist to undertake analysis of FRs in birds utilising landfill,

we recommend using internationally-standardised sampling protocols and eggs as the sampling

tissue of choice given their proven utility in environmental monitoring studies. Such studies

would benefit from the reporting of a minimum common suite of key chemicals. In terms of

PBDEs, a minimum of the following congeners should be measured: BDEs -28, -47, -99, -100, -

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153, -154, -183 and -209. Among HBCDDs, α-HBCDD, β- and γ-HBCDD should also be

quantified. It will be increasingly important to monitor other FRs given the restrictions applying

to legacy BFRs, the selection of which will depend on local legislation and use.

Although this review has reported elevated FR burdens in various landfill-associated birds,

we have also identified a piscivorous species that exhibits elevated concentrations of potentially

more toxic and bioavailable lower-brominated BDE congeners compared to a sympatric

congener species more reliant on terrestrial anthropogenic food (i.e., Audouin’s gull versus

yellow-legged gull, respectively, in the Mediterranean Sea). Further work is required to

understand the comparative importance of a piscivorous as opposed to a terrestrial anthropogenic

diet in terms of exposure potential to lower brominated congeners. Other research gaps include

elucidating the role of FR contamination in driving the declines seen in certain landfill-

associated birds, such as the herring gull and the glaucous gull, the relationship between bird use

of landfill and DDC-CO contamination (including potential effects on birds), and the impacts of

exposure to mixtures of anthropogenic chemicals on birds including those that use landfill.

Although birds rapidly metabolise many PFRs, this leads to underestimates of avian exposure to

these chemicals and further research in this area would also be desirable. Research is also

required to better understand and quantify the different exposure pathways of BFRs in landfill-

foraging birds.

Acknowledgments

The authors gratefully acknowledge provision of a CENTA studentship award to ADWT by the

UK Natural Environment Research Council (NERC ref. NE/L002493/1). This research also

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received funding from the European Union’s Horizon 2020 research and innovation programme

under the Marie Skłodowska-Curie grant agreement No 734522 (INTERWASTE) project.

Appendix A. Supplementary data

Supplementary data related to this article can be found at (web address)

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Table 1. Most recent (2001) industry-published figures for global ‘legacy’ brominated flame

retardant total market demand (in metric tonnes) by region. Source: BSEF (2003).

Legacy product

Americas Europe AsiaRest of World

World total

TBBPA 18000 11600 89400 600 119700

HBCDD 2800 9500 3900 500 16700

Deca–BDE 24500 7600 23000 1050 56100

Penta–BDE 7100 150 150 100 7500

Octa–BDE 1500 610 1500 180 3790

Total PBDEs 33100 8360 24650 1330 67440

Total BFRs 53900 29460 117950 2430 203740

Table 2. Concentrations of flame retardants measured in avian species associated with landfill in

various countries. Entries are listed by advancing date of publication of the study.

Species Sampling period

Location Tissues sampled (n = sample size)

Types of flame retardants detected

Reference

African sacred ibis(Threskiornis aethiopicus)

2004-2005

South Africa

Egg (n = 2) PBDEs, HBCDDs

(Polder et al., 2008)

White stork (Ciconia ciconia)

1999-2000

Spain Egg (n = 33)

NBFRs (Muñoz-Arnanz et al., 2010)

1999-2000

Spain Egg (n = 33)

PBDEs (Muñoz-Arnanz et al., 2011a)

1999-2000

Spain Egg (n = 33)

DDC-CO (Muñoz-Arnanz et al., 2011b)

Ring-billed gull(L.

2010 Canada Liver (n = PBDEs, (Gentes et al.,

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1315

1316

1317

1318

Species Sampling period

Location Tissues sampled (n = sample size)

Types of flame retardants detected

Reference

delawarensis) 28); plasma (n = 30)

NBFRs. DDC-CO

2012)

Herring gull (Larus argentatus)

2008 Canada Egg (n = ~200)

PBDEs (Chen et al., 2012a)

Yellow-legged gull(L. michahellis)

2010 Spain Egg (n = 36)

PBDEs (Morales et al., 2012)

Yellow-legged gull

2007 Spain Egg (n = 19)

DDC-CO (Muñoz-Arnanz et al., 2012)

Common starling(Sturnus vulgaris)

2009-2011

Canada Egg (n = 259)

PBDEs, NBFRs, DDC-CO

(Chen et al., 2013)

2009-2010

Canada Egg (n = 150)

PBDEs (Eens et al., 2013)

Ring-billed gull 2010–12 Canada Plasma (n = 76)

PBDEs (Gentes et al., 2015)

Common starling

2012 Canada Liver microsomes (n = 7)

PBDEs (Erratico et al., 2015)

Black kite (Milvus migrans)

2012-2014

Pakistan Tail feather (n = 8)

PBDEs, HBCDD, BTBPE

(Abbasi et al., 2016a)

Yellow-legged gull

2007 Spain Egg (n = 18)

PBDEs (Roscales et al., 2016)

Herring gull 1990–10 Canada Egg (n = ~1300)

PFRs (Greaves et al., 2016)

2014 USA Egg (n = 10)

DDC-CO (Su et al., 2017)

Cinereous vulture(Aegypius

2016 Spain Nestling down (n =

PBDEs, PFRs

(Monclús et al., 2018)

61

Species Sampling period

Location Tissues sampled (n = sample size)

Types of flame retardants detected

Reference

monachus) 16)

Black kite 2007-2016

Spain Egg (n = 51)

PBDEs, Dechloranes

(Blanco et al., 2018)

Table 3. A comparison of PBDE burdens reported in landfill-associated birds with reference

(i.e., non-landfill-associated) individuals in the literature.

Species Sampling

periodTissue Landfill

∑PBDEs

(n = sample size)

Reference ∑PBDEs

(n =sample size)

Source

White stork 1999–2000 Egg 2.79 – 20.5 (n = 10)*

0.214– 9.50 (n = 23)*

(Muñoz-Arnanz et al., 2011a)

Common starling

2009–2011 Egg 11–805 (n = 68)*

12–15 (n = 11)*

(Chen et al., 2013)

Ring-billed gull

2010–2012 Plasma 15.9 (n = 32)†

26.8 (n = 32)† (Gentes et al., 2015)

Black kite 2007–2016 Egg 1570 (n = 28) ‡

13.6 (n = 23) ‡

(Blanco et al., 2018)

Entries are listed by advancing date of publicationAll concentrations are in ng/g ww, except Blanco et al., 2018 (ng/g lw)* Range† Mean

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1326

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1328132913301331

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