Plastic Degradation and Its Environmental Implications with Special Reference to Poly(ethylene terephthalate).pdf

Mo 5,

Rich hi7,

Yuta 10,

Ma 13,

K 5

* Author for correspondence ([email protected]).

Electronic supplementary material is available at http://dx.doi.org/rstb20080284 or via http://rstb.royalsocietypublishing.org.

and human health.

Phil. Trans. R. Soc. B (2009) 364, 20272045

doi:10.1098/rstb.2008.0284

on April 14, 2015http://rstb.royalsocietypublishing.org/Downloaded from One contribution of 15 to a Theme Issue Plastics, the environment14National Institute of Public Health, Samsenthai road, Ban Kao-Gnod, Sisattanak District,Vientiane Municipality, LAO Peoples Democratic Republic

15Laboratory of Organic Geochemistry (LOG), Tokyo University of Agriculture and Technology, Fuchu,Tokyo 183-8509, Japan

Plastics debris in the marine environment, including resin pellets, fragments and microscopic plastic frag-ments, contain organic contaminants, including polychlorinated biphenyls (PCBs), polycyclic aromatichydrocarbons, petroleum hydrocarbons, organochlorine pesticides (2,20-bis(p-chlorophenyl)-1,1,1-tri-chloroethane, hexachlorinated hexanes), polybrominated diphenylethers, alkylphenols and bisphenolA, at concentrations from sub ng g1 to mg g1. Some of these compounds are added during plasticsmanufacture, while others adsorb from the surrounding seawater. Concentrations of hydrophobic con-taminants adsorbed on plastics showed distinct spatial variations reflecting global pollution patterns.Model calculations and experimental observations consistently show that polyethylene accumulatesmore organic contaminants than other plastics such as polypropylene and polyvinyl chloride. Both a math-ematical model using equilibrium partitioning and experimental data have demonstrated the transfer ofongsap Akkhavong14, Yuko Ogata15, Hisashi Hirai15, Satoru Iwasa15,

aoruko Mizukawa15, Yuki Hagino15, Ayako Imamura15, Mahua Saha1

and Hideshige Takada15,*1Marine Biology and Ecology Research Centre, Marine Institute University of Plymouth,

A403 Portland Square, Drake Circus, Plymouth, PL4 8AA, UK2School of Engineering and Electronics, University of Edinburgh, Old College, South Bridge Edinburgh

EH8 9YL, UK3Department of Civil, Construction and Environmental Engineering, North Carolina State University,

PO Box 7908, Raleigh, NC 27695, USA4Department of Water and Environmental Studies, Linkoping University, SE-581 83, Linkoping, Sweden5Marine Biology and Ecology Research Centre, Marine Institute, University of Plymouth, Drake Circus,

Plymouth PL4 8AA, UK6School of Biosciences, University of Exeter, Stocker Road, Exeter, EX4 4QD, UK

7Graduate School of Fisheries, Hokkaido University, Hakodate, Hokkaido 041-8611, Japan8Algalita Marine Research Foundation, 148 Marina Drive Long Beach, CA 90803, USA9Research Centre for Environmental Technology and Sustainable Development (CETASD),

Hanoi University of Science, Vietnam National University, T3 Building, 334 Nguyen Trai Street,Thanh Xuan District, Hanoi, Vietnam

10Economic, Social and Cultural Observation Unit, Office of the Council of Minister,Sahapoan Russi Blvd., Phnom Penh, Kingdom of Cambodia

11Science Education Department, De La Salle University, 2401 Taft Avenue, Malate,1004 Manila, The Philippines

12Environmental Research and Training Center, Bangkok, Technopolis, Klong 5, Klong Luang,Pathumthani 12120, Thailand

13Department of Environmental Sciences, Faculty of Environmental Studies, Universiti Putra Malaysia,43400 UPM, Serdang, Selangor Darul Ehsan, MalaysiaKTransport and release of chemicals fromplastics to the environment and to wildlifeEmma L. Teuten1,2, Jovita M. Saquing3, Detlef R. U. Knappe3,

rton A. Barlaz3, Susanne Jonsson4, Annika Bjorn4, Steven J. Rowland

ard C. Thompson1, Tamara S. Galloway6, Rei Yamashita7, Daisuke Oc

ka Watanuki7, Charles Moore8, Pham Hung Viet9, Touch Seang Tana

ricar Prudente11, Ruchaya Boonyatumanond12, Mohamad P. Zakaria2027 This journal is q 2009 The Royal Society

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4, sorption is described mathematically and the modelvalidatsummafoundenvironas a veSectiontic-derin 7, wof conFinallystrating

pore size in the polymer and the size of the additive

cleavage of the covalent bond(s) before migration canhemi-causelow).

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2028 E. L. Teuten et al. Chemicals from plastics to environment

Phil. Tra

on April 14, 2015http://rstb.royalsocietypublishing.org/Downloaded from ed by experimental observations. Section 5rizes the types and quantities of contaminantssorbed to plastics collected from the marinement. The remaining sections emphasize plastics

ctor in the transport of contaminants to animals.6 presents an overview of the transfer of plas-

ived contaminants to organisms. This is expandedhich describes literature concerning the transporttaminants to sediment-dwelling invertebrates., 8 reports initial experiments demon-

transfer of contaminants from plastics to

take place. Therefore, loss of reactively bonded ccals from the polymer resins is most probably beof release of unreacted constituents (see BPA be

In landfills, plastics are exposed to an extrsolvent in the form of acidic (pH 56) leawith high ionic strength and neutral or alleachates containing high-molecular-weight ocompounds. The different leachates have notdifferent potentials to extract and transportalso different biological populations with the potto degrade or transform the released additives.compounds. The following sections address the uptakeof contaminants from the environment onto plastics. In

and the surrounding environment. Release of a reac-tively bonded compound from a polymer requirescontaminants from plastic to organisms. A feedingcontaminated plastics to streaked shearwater chictutional monomers also present potential threatsfrom waste disposal sites into groundwater and/or suand polymers are complex phenomena dependentchemical properties of each additive. Bisphenol Adisposal sites in tropical Asia ranged from sub mgeconomic development.

Keywords: marine plastic debris; plasticendocrine-disrupting chemica

1. INTRODUCTIONPlastics are considered to be biochemically inert materialsthat do not interact with the endocrine system becauseof their large molecular size, which prohibits theirpenetration through the cell membrane. However,plastic debris present in the marine environment(marine plastics) carry chemicals of smaller molecularsize (MW, 1000). These chemicals can penetrate intocells, chemically interact with biologically importantmolecules and may disrupt the endocrine system. Suchchemicals are categorized into two groups: (i) hydro-phobic chemicals that are adsorbed from surroundingseawater owing to affinity of the chemicals for thehydrophobic surface of the plastics and (ii) additives,monomers and oligomers of the component moleculesof the plastics. Many of the contaminants addressedherein have known biological consequences. Forexample, the plastic constitutional monomer bisphenolA (BPA) and alkylphenol additives exert oestrogeniceffects (e.g. Sonnenschein & Soto 1998), while somephthalate plasticizers have been associated with reducedtestosterone production (e.g. Foster 2006). A wide rangeof biological effects have been reported for polychlori-nated biphenyls (PCBs; Neal 1985). Reviews of humanand wildlife exposure to plastics additives are alsoavailable in this volume (Koch & Calafat 2009; Meekeret al. 2009; Oehlmann et al. 2009).

The objective of this paper is to review the phenomenaby which plastics released to the environment serve ascarriers of organic contaminants to wildlife. The firsttwo sections describe leaching of contaminants fromplastics in landfills. Section 2 reviews the migration anddegradation of plasticizers (phthalates), additives(organotin compounds and nonylphenols (NP)) andmonomers (BPA), while 3 focuses on landfill leachateas a source of plastics-derived endocrine-disruptingns. R. Soc. B (2009)molecule are important parameters. Co-migrationand temperature are positive migration factors as arecertain physicalchemical properties of the additiveperiment indicated that PCBs could transfer from. Plasticizers, other plastics additives and consti-terrestrial environments because they can leachce waters. Leaching and degradation of plasticizersenvironmental conditions in the landfill and the

oncentrations in leachates from municipal waste1 to mg l1 and were correlated with the level of

in pellet; microplastics; landfill leachate;persistent organic pollutants

higher-trophic-level organisms (acronyms in this paperare listed in table 1).

2. RELEASE AND DEGRADATION OFADDITIVES AND CONSTITUTIONALMONOMERS FROM POLYMERSOrganic compounds are used as additives in polymers toimprove the properties of the resulting products. Releaseof the additives to the surrounding environment is anunwanted process for both the manufacturer and theenvironment, since loss of additives shortens polymer life-time, e.g. loss of plasticizers lowers the tensile strength ofpolyvinyl chloride (PVC; Boyer 1951), and living organ-isms are exposed to the released additives. Phthalates,organotins and BPA, mentioned subsequently, havebeen shown to target nuclear hormone receptor signallingpathways (Grun & Blumberg 2007). The release may takeplace during the service life of the plastics or after theirdisposal, for example in landfills. Both the landfill com-partment and other potential receptors such as sedimentsrepresent complex environments with multiple chemicaland biological processes occurring concurrently.

The migration potential of an additive in a polymerdepends on several parameters. The polymer itselfhas a three-dimensional porous structure in whichthe additives are dispersed. The pore diameter andthe size of the additive are correlated such that smaller(lower molecular weight) additives move more easilythrough a polymer with bigger pore size. Additivesthat fit more exactly in the pores have a small butnot insignificant capacity to migrate. Therefore, the

Chemicals from plastics to environment E. L. Teuten et al. 2029

on April 14, 2015http://rstb.royalsocietypublishing.org/Downloaded from Table 1. List of acronyms.

acronym meaning

BD brominated diphenylether congenerBDEs brominated diphenylethers

BPA bisphenol ACB chlorinated biphenyl congenerDDD 2,20-bis(p-chlorophenyl)-1,1-dischloroethaneDDE 2,20-bis(p-chlorophenyl)-1,1-dischloroethyleneDDT 2,20-bis(p-chlorophenyl)-1,1-trichloroethaneDDTs DDT and its metabolites (i.e. DDD and DDE)DEHP diethylhexyl phthalateDMP dimethyl phthalateDOC dissolved organic carbon

E1 estroneE2 oestradiolE3 estriolEDCs endocrine-disrupting chemicalsEEQ oestradiol-equivalent concentration

FTIR Fourier transform infrared spectroscopyGC-ECD gas chromatograph equipped with an electron

capture detectorGDP gross domestic productsHCHs hexachlorocyclohexanesPlasticizers, which are the largest group of additivesin polymers, range from molecular weights of approxi-mately 200 to almost 700 g mol1 and cover watersolubility from g l1 to sub-mg l1. Phthalates (ormore chemically correct, alkyl/aryl esters of 1,2-benze-nedicarboxylic acid) are the most common plasticizersand may account for more than 60 per cent of poly-mers of PVC (Giam et al. 1984). Dimethyl phthalate(DMP) is fairly easily released from its resin, as soonas the DMP-containing product is landfilled, owingto its relatively high water solubility, i.e. there is a con-tinuous depletion of DMP from the resin surface, andthe negative concentration gradient from the inside tothe surface causes the migration. In contrast, thehigher-molecular-weight phthalates, such as diethyl-hexyl phthalate (DEHP), are more resistant tomigration owing to their hydrophobicity, whichcauses less release from the polymer surface to leachatecompared with DMP.

The importance of the surrounding medium for theextraction potential can be exemplified by the differentdegradation phases in a landfill. Acidic pH and highionic strength of the leachate that surrounds waste

HDPE high-density polyethyleneHOCs hydrophobic organic contaminantsMW molecular weightNOEC no-effect concentrationNP nonylphenol

OP octylphenolPAHs polycyclic aromatic hydrocarbonsPBDEs polybrominated diphenylethersPCBs polychlorinated biphenylsPCE tetrachloroethylene

PE polyethylenePVC polyvinyl chlorideSML sea-surface microlayerSOM sorbent organic matter

Tg glass transition temperatureTNP trisnonylphenolphosphitesUV ultraviolet

Phil. Trans. R. Soc. B (2009)materials lower the release potential of organic com-pounds, which make the initial acidogenic phase in alandfills development a very poor extraction solventfor water-resistant plasticizers (Bauer et al. 1998;figure 1). In contrast, a neutral leachate, as found inlandfills in the stable methanogenic phase, containingcolloidal humic material, facilitates leaching and trans-port of non-soluble plasticizers owing to sorption tothe dissolved organic carbon (DOC) fraction.Therefore, concentrations of phthalate esters in landfillleachates are highly correlated to the DOC content(Bauer & Herrmann 1998). As a consequence of thedepletion of plasticizer from the polymer surface,migration from the inner part of the polymer productis enhanced. However, migration from the inner partto the outer surface seems to slow down and evenstop as the polymer reaches its glass transition state(Ejlertsson et al. 2003). Then, new release of plastici-zers only occurs if the brittle polymer structurefractures to expose new surfaces.

Degradation of phthalates is initiated by hydrolysis ofthe ester moiety to phthalic acid and the correspondingalcohols via the monoesters. In landfills, biotic hydroly-sis is far more important than abiotic hydrolysis

hydrophilic moderate hydrophobic hydrophobic

time

conce

ntr

atio

n

Figure 1. Schematic appearance and concentration of ahydrophilic (left), moderate hydrophobic (middle) and hydro-phobic (right) phthalic acid diester (solid lines) and respectivemonoester (dashed lines) in landfill leachate (modified fromJonsson 2003). The appearance of the diester is correlated

to its depletion in the phthalate-containing product.(Furtmann 1996; Staples et al. 1997) and takes place(i) at the surface of the original products, (ii) afterthey have been released from the products and dissolvedin the leachate or (iii) following release from anothersurface to which they adsorbed after leaving the originalresin. The most important hydrolysis scenario dependson the water solubility of the phthalate, i.e. the solublephthalates are probably hydrolysed in the water phaseand the hydrophobic phthalates are hydrolysed on tosolid surfaces. Hydrolysis is strongly correlated to themethanogenic flora ( Jonsson et al. 2003a, 2006;figure 2). Accumulation of the monoester occurs ifthe hydrolysis rate of the diester to the monoester isfaster than that of the monoester to phthalic acid(Vavilin et al. 2005). In fact, phthalate monoester con-centrations have been observed at higher concentrationsthan the corresponding diesters in landfill leachates( Jonsson et al. 2003b). If the phthalate diester isslowly released during a longer period includingthe methanogenic stage, the time period when the

under landfill conditions has not been reported as far


on April 14, 2015http://rstb.royalsocietypublishing.org/Downloaded from monoester is observed in the leachate is prolonged andthe concentrations of the diester are consequentlylower.

Organotin compounds are used as stabilizing additivesin polymers, such as PVC, and they deserve specialattention because of their toxicity such as deteriorationof human immune function and endocrine disruption(Batt 2006). The stabilizers are added as high molecu-lar mono- and dialkyltin carboxylates, mercaptides upto 0.54 per cent or, more common, mercaptans orsulphides up to 0.18 per cent, calculated as tin, in thepolymer (Murphy et al. 2000; Batt 2006). The carbox-ylates and mercaptides are rapidly hydrolysed to theirmono- and dialkyltin species, respectively, when incontact with water (Bjorn 2007). The alkyltins arealso hydrolysed when they act as stabilizers withinthe polymer and are consequently released from thepolymer surface as alkyltin chlorides. As for the phtha-lates, it seems likely that the main release of organotincompounds from plastic material occurs when a land-fill turns methanogenic (Bjorn et al. 2007). It has beenshown that the tin stabilizers are co-extracted from thepolymer together with the phthalates. Therefore, orga-notins in flexible polymers are more readily releasedthan from rigid ones. It should be noted that 90 percent of the tin stabilizers are used in rigid PVC

initial methanogenic phase

acidogenic phase stable methanogenic

methane

conce

ntr

atio

n

Figure 2. Degradation of a phthalic acid diester (solid line) toits corresponding monoester (dashed line) and phthalic acid(dotted line) in a landfill developing from acidogenic tostable methanogenic phase. Also, the methane production is

included (reproduced with permission from Jonsson 2003).(ESPA 2002). However, at temperatures above theglass transition of the polymer, more organotin com-pounds are released than at temperatures below thispoint (Bjorn et al. 2007).

The alkyltin compounds may dealkylate to inor-ganic tin, methylate or demethylate in the landfillenvironment. It is likely that the microbial methylationcapacity is greater at higher concentrations (more than500 mg Sn l1), while demethylation occurs at lowertin concentrations (below 100 mg Sn l1; Bjorn2007). Formation of tetramethyl tin changes the prop-erties of the tin species radically, since this compoundis very volatile.

Alkylphenols can be used as plasticizing additives oras stabilizers when added as derivatives of phosphites(e.g. trisnonylphenolphosphites: TNP). Upon oxi-dation and hydrolysis, alkylphenol phosphites arehydrolysed to the corresponding alkylphenol and phos-phate, for example, TNP is readily oxidized and

Phil. Trans. R. Soc. B (2009)as we know, but results from an analysis of landfillleachates suggest that the additive or unreacted BPA,owing to its more hydrophilic character, is readilyreleased from its polymer during the early age of alandfill (Asakura et al. 2004, i.e. under acidogenic con-ditions as for the phthalate DMP). This is supportedby leaching studies with water-containing acetic acidand ethanol (Kawamura et al. 1998), which is expectedto mimic acidogenic leachates. Concerning degra-dation, complete mineralization has only beenreported under aerobic conditions (e.g. Zhang et al.2007). Bisphenol A is reported to be preserved inanaerobic sediments (e.g. Ying & Kookana 2003).

3. PHENOLIC ENDOCRINE-DISRUPTINGCHEMICALS IN LEACHATES FROMMUNICIPAL WASTEConsiderable amounts of plastics are disposed of inmunicipal landfills. As indicated earlier, certain addi-tives and monomers can be released from plastic andwill consequently be present in landfill leachate.Detection of BPA, phthalates and the alkylphenolsNP and octylphenol (OP) in landfill leachate has beenreported (Yasuhara et al. 1997; Yamamoto et al. 2001;Fromme et al. 2002; Coors et al. 2003; Jonsson et al.2003c; Asakura et al. 2004; Deng et al. 2006).Although BPA concentrations varied depending onwaste composition and landfill operation, concen-trations of BPA in leachates ranged from ten to tenthousand mg l1 from sites in the USA (Coors et al.2003), Germany (Fromme et al. 2002) and Japan(Asakura et al. 2004). These concentrations weremuch higher than those detected in municipal sewageeffluents (approx. 0.010.1 mg l1, Fromme et al.2002; Nakada et al. 2004), implying that untreated lea-chates from the landfills are potentially significantsources of BPA for the aquatic environment.Furthermore, the BPA concentrations in the leachateshydrolysed to NP under ambient conditions (Murata1999). Since the alkylphenols and the phosphites areadditives, the same reasoning can be applied for thesecompounds as for the phthalates. More precisely, com-pounds with shorter alkyl chains have higher leachingpotential than longer alkyl chain analogues, and metha-nogenic leachates are more extractive than acidogenicleachates. However, unlike the phthalates, alkylphenolphosphites are only used in concentrations up to 3per cent, compared with 60 per cent for the phthalates.Degradation studies of the pure alkylphenols performedunder landfill conditions are scarce. However, alkyl-phenols seem to be the ultimate degradation productwhen alkylphenol ethoxylates are transformed undermethanogenic conditions (e.g. Ejlertsson et al. 1999),suggesting that no further degradation occurs underanaerobic conditions (Maguire 1999).

Bisphenol A is, in contrast to the aforementionedcompounds, mainly used as a building block ofpolycarbonate plastics, where the alkylphenol p-tert-butylphenol is added as a polymerization adjustor, oras a key constituent together with epichlorohydrin ofepoxy resins. Also, BPA is used as additive in PVC,printer ink and some other products. Release of BPA

hen

(ED


on April 14, 2015http://rstb.royalsocietypublishing.org/Downloaded from were up to seven orders of magnitude higher than theno-effect concentration (NOEC) of BPA for endocrinedisruption in freshwater organisms (i.e. at 8 ng l1 toinduce malformations in female organs of a freshwatersnail, Marisa cornuarietis; Schulte-Oehlmann et al.2001). Significant concentrations of NP were alsodetected in landfill leachate sites (Asakura et al.2004). However, the reported concentration ranges ofNP (Asakura et al. 2004) were similar to those inmunicipal sewage effluents (Nakada et al. 2004).

Economic growth and industrialization bring largeramounts of plastics into society and may increase theamount of plastic waste. To investigate the effect ofindustrialization on the presence of endocrine-disrupting chemicals (EDCs) in landfill leachates, wemeasured plastic-derived chemicals in leachates fromtropical Asian countries at different stages of economicgrowth. Leachate samples were collected from opendumps in Malaysia (Kuala Lumpur), Thailand(Bangkok), The Philippines (Manila), Vietnam(Hanoi, Can Tho), Cambodia (Phnom Penh, Angkor),Laos (Vientiane) and India (Kolkata) between 2002and 2006. At all the sites, municipal wastes, includingplastics, are buried. As a reference, leachate samplescollected from a landfill site in Japan were also collectedand analysed for the EDCs. Details of the analytical

nonylphenol

10 000

1000

100

10

1

0.1

0.01

0.001octylphenol bisp

conce

ntr

atio

n (

g l1

)

Figure 3. Concentrations of endocrine-disrupting chemicalsprocedure were described by Nakada et al. (2004, 2006).Concentrations of the EDCs in the leachates from

sites in tropical Asia and Japan are shown in figure 3.Among the EDCs measured, BPA showed highestconcentrations in the tropical Asian leachates, rangingfrom 0.18 to 4300 mg l1. The highest concentrationswere observed in leachates from Malaysia and werecomparable to those from Japan. The concentrationrange of NP in the leachates (0.1898 mg l1) waslower than BPA but higher than OP (0.033.4 mg l1).Bisphenol A concentrations were one to five ordersof magnitude higher than those in sewage effluents(Nakada et al. 2004), whereas NP concentrations inthe leachates were one to two orders of magnitudehigher than those in effluents. This highlights theimportance of the leachate as a source of BPA inaquatic environments. Bisphenol A in leachate couldbe derived from unreacted monomers in disposedpolymers (polycarbonates and epoxy resins),

Phil. Trans. R. Soc. B (2009)degradation of the polymers and additives. In manylandfill sites in industrialized countries, treatmentfacilities are installed and the environmental burdenof these EDCs is reduced. High removal efficiency ofBPA has been reported with aerobic treatment(99.399.7%, Kawagoshi et al. 2003; Asakura et al.2004) and with membrane bioreactors (95.3%,Wintgens et al. 2003). However, because of highconcentrations of BPA in raw leachates, even treatedleachates showed higher BPA concentrations (0.1130 mg l1, Wintgens et al. 2003; Asakura et al. 2004)than the NOEC to freshwater organisms (0.008 mg l1).Bisphenol A is more problematic in tropical Asianlandfill sites with either no, or poorly functioning,leachate treatment facilities. Consequently, high con-centrations of BPA were discharged to the surroundingenvironment (e.g. rivers, groundwater). Notably, BPAconcentrations in water samples from a Malaysianpond, into which the leachate from the dump flowed,were an order of magnitude higher (i.e. approx.11 mg l1) than in the upstream inflowing river(0.45 mg l1). This clearly demonstrates that waste-plastic-derived chemicals significantly increase theconcentrations of EDCs in the environment.

High concentrations of natural oestrogens (estrone,E1: 0.1271.00 mg l1; oestradiol, E2: 0.002

ol A estrone oestradiol oestriol

LaosCambodiaVietnamIndiaThailandThe PhilippinesMalaysiaJapan

Cs) in leachates from waste disposal sites in Asia.0.0243 mg l1) were also detected in the leachatefrom the tropical Asian countries. This is in contrastto leachate from a Japanese landfill site where relativelylow concentrations of natural oestrogens (E1:,0.05 mg l1; E2: ,0.008 mg l1) were detected.The natural oestrogens in the landfill leachates fromthe Southeast Asian countries could be derived fromthe disposal of human wastes and/or input from thefaeces of scavengers living at the dumping sites. Basedon the concentrations of individual EDCs and relativepotency of endocrine disruption for individual com-pounds, oestrogenic activities of the individualcompounds have been calculated and compared. Thefollowing relative potencies reported by Sumpter &Johnson (2005) for fish were used: NP (0.0025), OP(0.002), BPA (0.0004), E1 (0.3) and E2 (1.00).Oestradiol-equivalent concentrations (EEQ) werecalculated by multiplying the concentrations of the indi-vidual compounds by their relative potency. The total

which is characterized on the basis of its internal struc-ture as either glassy or rubbery. Hence, SOM can beenvisioned as a mixture of glassy and rubbery poly-mers. The polymer segments of the glassy phaseshave higher cohesive forces and are more condensed,whereas those of the rubbery phases exhibit greatermobility and flexibility and can be visualized as adynamic viscous liquid (Tobolsky & Mark 1980).A particular polymer can transition from the rubberyto the glassy state when it is cooled below its glasstransition temperature (Tg). Based on Tg, a polymer isclassified as rubbery or glassy at a given environmentaltemperature. At room temperature, polymers that havea low Tg (e.g. polyethylene has a Tg of2688C) are rub-bery, while those that have a high Tg (e.g. PVC has aTg of 808C) are glassy (Brandup et al. 1989). The poly-mer characteristics of the crystalline region lie betweenthose of rubbery and glassy polymers owing to theirunique structure. Crystalline polymers are moderatelyhard, yet flexible and strong (Treloar 1974).

Glassy polymers, because of their rigidity, possesslong-lived, closed internal nanoscale pores that canserve as adsorption sites. The existence of amorphouspolymer segments and internal nanovoids in glassypolymers results in HOC sorption by linear dissolution(partitioning/absorption) and by nonlinear hole-filling


on April 14, 2015http://rstb.royalsocietypublishing.org/Downloaded from 4. SORPTION AND DESORPTION OFANTHROPOGENIC CONTAMINANTSFROM PLASTICSSorption and desorption are essential fate processesgoverning the distribution, persistence and ecologicalimpact of hydrophobic anthropogenic contaminants interrestrial and aquatic systems. Anthropogenic contami-nants such as alkylbenzenes, chlorinated hydrocarbons,polycyclic aromatic hydrocarbons (PAHs) and PCBsare examples of compounds that will probably associatewith sorbent organic matter (SOM) in the environment.The association of hydrophobic organic contaminants(HOCs) with SOM retards their transport and reducestheir availability for biological and chemical transform-ation. Traditionally, the organic fraction of soils andsediments was considered to be the most importantform of SOM in the environment, but recent studiesdocumented the importance of plastics in sedimentsand debris collected from the marine environment(Colton et al. 1974; Mato et al. 2001; Ng & Obbard2006; Rios et al. 2007). Hydrophobic organic contami-nants were shown to have greater affinity for a range ofplastics (polyethylene, polypropylene, PVC) comparedwith natural sediments (Teuten et al. 2007) and weredetected on plastic pellets collected from the marineenvironment (Mato et al. 2001; Rios et al. 2007) asdescribed in 5.

The extent and rate of HOC (de)sorption areinfluenced by factors including sorbent (i.e. SOM) prop-erties, sorbate (i.e. HOC) properties, dissolved organiccompounds in the aqueous phase, pH and temperature.EEQ ranged from 3.4 to 1355 ng-E2 l1. The highestEEQs were observed in Malaysian, The Philippinesand Thai leachates, where much higher concentrationsof BPA were observed. In those leachates, BPAaccounted for over 50 per cent of the total EEQ. Thishighlights the importance of BPA in terms of endocrinedisruption caused by leachate. The abundance of BPAover natural oestrogens in the leachate contrasts withmunicipal wastewater effluents where natural oestro-gens usually dominate over synthetic chemicals (e.g.Desbrow et al. 1998; Nakada et al. 2004).

Among the countries investigated, the more indus-trialized countries (e.g. Malaysia and Thailand) hadhigher BPA concentrations in landfill leachate thanless industrialized countries (e.g. Laos and Cambodia).To quantitatively express this trend, BPA concen-trations were plotted against per capita gross domesticproducts (GDPs; Earth Trends 2007) in figure 4.Bisphenol A concentrations in the leachate show asignificant positive correlation with per capitaGDP of the tropical Asian countries (r2 0.66, n 26,p, 0.0001). The most probable reason is that moreindustrialized countries use larger quantities of plasticsresulting in the generation of more plastic waste. Thissuggests that economic growth in developing countriesmay increase the environmental prevalence of EDCsunless the leachate is collected and properly treated.To reduce the input of EDCs to the environment, theamount of waste plastics discarded should be decreasedthrough reduction, recycling or other methods of disposalof plastic.Phil. Trans. R. Soc. B (2009)The following discussion will focus on the effects ofsorbent properties on sorption equilibrium and(de)sorption kinetics.

Sorbent organic matter in the environment is com-posed of organic polymers that contain crystalline andamorphous regions. The crystalline region is charac-terized by molecules or segments of molecules thatare regularly arranged in a crystal lattice. In contrast,the amorphous region has randomly arranged mole-cules, thus exhibiting a structure that is loose andflexible, and more similar to liquids. Sorption ofHOCs generally occurs in the amorphous region,

10 000

1000

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10

1

0.11000 10 000 30 000

per capita GDP (USD)

bisp

heno

l A in

leac

hate

(m

L1 )

Figure 4. Relationship between BPA concentrations inleachates from waste disposal sites and per capita GDP ofAsian countries (r2 0.66, n 26). Leachates from municipalwaste disposal sites in capital and other major cities are plotted.See figure 3 for the symbols of the countries. Data are for

countries except Japan.

and each sorbent particle is a homogeneous polymericsphere. Ficks second law of diffusion can be used toexpress HOC diffusion from plastic particles. In


on April 14, 2015http://rstb.royalsocietypublishing.org/Downloaded from radial coordinates, Ficks second law of diffusionyields equation (4.1):

@q

@t D

r2@

@rr2

@q

@t

; 4:1

where D is the diffusion coefficient (L2/T), q the solid-phase concentration (sorbed HOC mass/sorbent mass),r the radial position in the sorbent particle (L) and t thetime.

To solve equation (4.1), two-dimensionless variables,T and R, are introduced:

T Dta2

and R ra

, 4:2

where a is the sorbent particle radius.Therefore, the governing equation is transformed to:

@q

@T 1

R2@

@RR2

@q

@R

: 4:3

Initial and boundary conditions specific to theexperimental method employed to estimate D arerequired to solve equation (4.3). For the initial con-dition, it was assumed that sorption equilibrium wasattained prior to initiation of desorption and, there-fore, the solid-phase concentration (q0) was uniformthroughout the sorbent at the beginning of desorption(adsorption) mechanisms (Xing & Pignatello 1997).Because of the dual sorption mechanisms, the sorptivebehaviour of glassy polymers is normally described bythe nonlinear Freundlich model (q Kf Cne ), where q isthe amount of the compound sorbed per unit mass ofsolid, Ce the aqueous-phase concentration at equili-brium, Kf the Freundlich constant related to thecapacity of the sorbent material to sorb the sorbateand n the Freundlich exponent and an indicator ofthe site energy distribution of a sorbent (i.e. sorbentheterogeneity increases as n decreases from 1; Carteret al. 1995). Absorptive partitioning into an organicmatrix is characterized by a linear sorption model(q KpCe), where Kp is the partition coefficient.Weber et al. (1992) showed that nonlinear behaviourmay be masked at high aqueous-phase concentrations,but can actually control the overall sorption behaviourat low aqueous-phase concentrations. At low-phaseconcentrations (,11.5% of aqueous solubility),HOCs are sorbed most favourably by regions or com-ponents of SOM that have the strongest affinity forthat compound (Chiou & Kile 1998). As the high-affinity regions (characterized by nonlinear sorptionisotherms) become saturated, HOC sorption is limitedto less strongly sorbing regions (characterized by linearsorption isotherms).

(a) Model descriptionDesorption of HOCs from plastics can be describedby a one-compartment polymer diffusion model.The model assumes that the HOC desorption rate islimited by diffusion through a single polymer phase,Phil. Trans. R. Soc. B (2009)tests, i.e.

q q0 at T 0 from 0 R 1: 4:4The first boundary condition requires that sym-

metry is maintained at the particle centre at alltimes, i.e.

@q

@R 0 at R 0 and T 0: 4:5

The second boundary condition specifies the solid-phase concentration at the external solid surface. Forthe results described here, in which volatile HOCswere tested, sorbents equilibrated with an aqueousphase were sparged continuously during the desorptiontest. Thus, the aqueous-phase HOC concentration wasnegligible (i.e. an infinite sink was approximated).

Assuming instantaneous equilibrium between thesolid- and aqueous-phase concentrations at the externalsorbent surface, the solid-phase concentration at theexternal particle boundary was therefore also zero, i.e.

q 0 at R 1: 4:6A CrankNicholson finite-difference algorithm was

developed to solve the one-compartment polymer dif-fusion model. The NewtonRaphson optimizationroutine was used to determine the diffusion coefficient(D) such that the mean square error between themodel output and experimental data was minimized.The model requires the following input parameters:isotherm parameters (Kp for linear isotherms and Kfand n for nonlinear isotherms), particle radius, particledensity, fractional uptake and initial estimate of D.

(b) Model applicationThe validity of the one-compartment polymer diffu-sion model to simulate desorption kinetics of HOCsin homogeneous plastics was tested using toluene,o-xylene and tetrachloroethylene (PCE) as modelHOCs, high-density polyethylene (HDPE) as amodel rubbery polymer and PVC as a model glassypolymer. As shown in figure 5, model results agreedwell with o-xylene desorption data from HDPE andPVC. Similarly, model results agreed well with toluene(figure 6) and PCE (not shown) desorption data.Table 2 summarizes inputs to the desorption modeland the estimated D values for the tested HOCs.

In general, HOC diffusivities in plastics were higherin HDPE (D 10210 cm2 s1) and lower in PVC(D 1021310214 cm2 s1). The diffusivity of HOCsin PVC is reasonably consistent with the values observedby Berens (1989). For HDPE, the diffusivity of HOCs isan order of magnitude lower than values reported inliterature, possibly because of differences in polymercomposition and crystallinity, experimental conditionsand uncertainties in estimates of diffusional length(Roger et al. 1960; Park et al. 1996; Sangam &Rowe, 2001; Joo et al. 2004). Typically, HDPE has8095 per cent crystallinity (Brandup et al. 1989) butthe crystallinity of an HDPE geomembrane tested bySangam & Rowe (2001) was only 47 per cent, whichcould account for the difference in D estimates.Moreover, uncertainties in diffusional length scale (i.e.film thickness or particle radius) could also affect the

ti14


on April 14, 2015http://rstb.royalsocietypublishing.org/Downloaded from 0

norm

aliz

ed m

ass r

emai

ning

(%)

121086420

102030405060708090

100calculated diffusion coefficients. When the diffusioncoefficients are normalized by diffusional length scale(D/a2), the normalized HOC diffusion coefficients deter-mined for HDPE in this study have the same order ofmagnitude (1022 d1) as those in previous reports(Sangam & Rowe 2001; Joo et al. 2004).

The difference in HOC desorption rates observedbetween HDPE and PVC is consistent with their rub-bery and glassy states. The polymeric organic matrix ofglassy polymers such as PVC is more rigid than that ofrubbery polymers such as HDPE. Because the relax-ation speeds of glassy polymers are slow, diffusion ofsolute molecules into and out of the condensed andhighly cross-linked organic matter is slow, whichexplains the smaller HOC diffusivities in glassy poly-mers (Brusseau et al. 1991; Pignatello & Xing 1996;

PVC 0.14 mmmodel PVC 1.7 mmmodel HDPE 0.5 mmmodel HDPE 2 mm

Figure 5. Comparison of o-xylene desorption data and one-compdesorption rates from PVC and HDPE spheres of different diame

libration in ultrapure water.

100908070605040302010

PVC toluene mass remainingHDPE toluene mass remainingPVC toluene mass releasedHDPE toluene mass released

0

1009080706050403020100

0 1 2 3time (d)

mas

s of t

olue

ne re

leas

ed (

g g

1 )

sorb

ed m

ass o

f tol

uene

rem

aini

ng (

g g

1 )

4 5 6

Figure 6. Effect of polymer type on sorbed toluene massremaining and released per gram of sorbent. Lines representmodel fits for sorption equilibrium liquid-phase concen-

tration (Ce) of 100 mg l1, Kp 70.7 (mg kg1)(l mg1) for

HDPE and Kf 1663 (mg kg1)(l mg1), and n 0.864for PVC (Wu et al. 2001). The particle diameters ofHDPE and PVC were 0.5 and 0.14 mm, respectively.

Phil. Trans. R. Soc. B (2009)Huang & Weber 1997, 1998). Moreover, nanovoidswithin glassy polymer matrices provide strong adsorp-tion sites, and desorption of HOCs from these sites isgenerally activated (Pignatello & Xing 1996).

For the same input parameters and estimate of D,figure 5 illustrates the effect of diffusional length scaleon o-xylene desorption rates. The particle sizes, forwhich model predictions are shown in figure 5, arerepresentative of the size range of plastic pellets andfragments collected from the marine environment(Colton et al. 1974; Mato et al. 2001; Rios et al.2007). In agreement with the inverse proportionalitybetween desorption rate and the square of the sorbentparticle radius, the results in figure 5 illustrate that

323028262422201816me (d)

HDPE 0.5 mmmodel PVC 0.14 mm

model HDPE 1 mmmodel HDPE 5 mm

artment diffusion model fits as well as predictions of o-xyleneters. Desorption data were measured after six months of equi-HOC desorption rates decreased dramatically as thediffusional length scale increased. Additional modelpredictions showed that the time required for 50 percent desorption of toluene, o-xylene and PCE from rub-bery plastics (e.g. polyethylene and polypropylene) witha 1 mm particle diameter was 2.8, 4.0 and 6.2 days,respectively. When the particle diameter was doubled(2 mm), the half-lives increased to 11.3, 16.1 and25.3 days, respectively. For glassy plastics (e.g. PVCand polystyrene) with a particle diameter of 0.2 mm,the time required for 50 per cent desorption of toluene,o-xylene and PCE was 85 days, 2.3 years and 6.5 years,respectively. The predicted half-life of o-xylene andPCE in PVC was.100 years when the particle diameterof glassy polymers was increased to 1.7 mm.

Figure 6 compares the mass of toluene releasedper unit mass of HDPE and PVC, assumingCe 100 mg l1 and using a particle diameter of0.5 mm for HDPE and 0.14 mm for PVC. Althoughtoluene diffuses three orders of magnitude faster inHDPE, the amount of toluene released from PVC isgreater than that released from HDPE. This is becausethe mass of toluene sorbed to PVC at equilibrium isgreater than the mass of toluene sorbed to HDPE(Wu et al. 2001).

(2001) triggered a series of systematic studies on toxic

xyl

8


on April 14, 2015http://rstb.royalsocietypublishing.org/Downloaded from chemicals in marine plastics. They detected PCBs inpolypropylene pellets from Japanese coasts with concen-trations ranging from 4 to 117 ng g1. They conducted afield adsorption experiment using virgin polypropylenepellets and demonstrated a significant and consistentincrease in PCB concentrations throughout the 6 dayexperiment. This indicated that the source of thePCBs was ambient seawater and that adsorption to thepellet surfaces was the mechanism of enrichment. Inanother adsorption experiment, Mato et al. (2002)In summary, results from alkylbenzene and PCEdesorption kinetic tests for glassy and rubbery poly-mers suggest that both sorbent and sorbate propertiesstrongly influence HOC sorption uptake and desorp-tion kinetics. Glassy polymers exhibit larger HOCsorption capacities and slower HOC release rates thanrubbery polymers. Moreover, the size of the plasticpellet or fragment strongly affects the rate at whichsorbed HOCs are released.

5. TYPES OF CONTAMINANTS DETECTED INMARINE PLASTICS(a) Adsorption of contaminants to marineplastics from surrounding seawater

Carpenter et al. (1972) first detected PCBs inpolystyrene spherules collected from Niantic Bay (north-eastern Long Island Sound, USA). Although theysuggested that the PCBs were adsorbed onto the plasticfrom the surrounding seawater, no supporting evidencewas provided. After a 30 years break, Mato et al.

Table 2. Model input parameters and estimates of toluene, o-

material

particledensity(g cm3)

meanparticlediameter(mm)

isotherm parameters

tolueneb o-xyleneb

K nd K nd

HDPE 0.962 500 70.7e 1.0 244.1 1.0PVC 1.4 140 1663f 0.864 4634 0.71

acm2 s1.bValues from Wu et al. (2001).cValues from Wagner (2003).dDimensionless Freundlich exponent.eKp for HDPE (mg kg

1)(l mg1).fKf for PVC (mg kg

1)(l mg1)n.subjected polyethylene and polypropylene pellets to sea-water for two weeks and found that polyethylene pelletsadsorbed four times more PCBs than polypropylenepellets, indicating that polyethylene has higher affinityfor hydrophobic compounds. This is consistent withfield observation and experimental work described laterand literature (e.g. Karapanagioti & Klontza 2008).Comparison of PCB concentrations in marine plasticresin pellets with those in seawater suggested their highdegree of accumulation (apparent adsorption coefficientof 105106).

Subsequently, Endo et al. (2005) conducted adetailed study of PCBs in beached resin pellets. Theyanalysed PCB concentrations in individual pellets andobserved a large (i.e. two orders of magnitude) piece-to-piece variation in PCB concentrations among the

Phil. Trans. R. Soc. B (2009)pellets. Variation in PCB concentrations may becaused by various factors including difference inmaterials, weathering and residence time in the sea.Endo et al. (2005) found that polyethylene pelletstended to have higher concentrations of PCBs thanpolypropylene, consistent with the results of the aboveadsorption experiment. They also indicated that yellowingand/or fouled pellets had higher concentrations of PCBs.Yellowing is derived from oxidation of phenolic additivesto quinone-type degradation products and, therefore, isan indication of environmental residence time of thepellets. Based on their findings, they proposed a monitor-ing methodology pellet watch where beached yellowedand/or fouled polyethylene pellets are used to monitorcoastal pollution by hydrophobic chemicals.

Based on the results, the monitoring was expandedto a global scale, named International Pellet Watch(http://www.tuat.ac.jp/~gaia/ipw/index.html). In thisscheme, plastic resin pellets are collected on beachesby local volunteers and sent to the Tokyo Universityof Agriculture and Technology for analysis of a varietyof HOCs. Through the activities of volunteers world-wide, and a network of scientists, 27 samples from16 countries have been analysed. As shown infigure 7, PCB concentrations of beached plastic resinpellets were highest on the coasts of USA(San Francisco, Los Angeles and Boston), followedby Japan and Europe (The Netherlands, UK andItaly). In tropical Asia, Australia and southernAfrica, PCB concentrations were much lower. Thisregional pattern reflects a difference in PCB usage,with larger amounts of PCBs used in the USA, western

ene and PCE diffusivities in HDPE and PVC.

diffusion coefficienta

PCEc toluene o-xylene PCE

K nd D D D

385 1.0 3.13 1010 2.20 1010 1.42 10102951 0.918 4.33 1013 4.22 1014 1.50 1014Europe and Japan, and minimal usage in tropical Asia,Australia and southern Africa. For example, more thanhalf of the total global production of PCBs was used inthe USA (Erickson 1997). Discharged PCBs haveaccumulated in coastal zones, particularly in sedi-ments, which are likely to be resuspended into thewater column. Correspondingly, higher concentrationsof PCBs are still found in sediments, seawater andmarine biota in the USA, western Europe and Japan.To examine the feasibility of a monitoring methodologyemploying plastic pellets, concentrations of PCBs inthe pellets were compared with those in conventionalbiomonitoring organisms, i.e. mussels (green mussels,Perna viridis; blue mussels, Mytilus galloprovincialis) inthe corresponding zones. Polychlorinated biphenylconcentrations in the beached resin pellets were highly

0.1

PCBsDDEPAHs

PBDEs

NP

OPBPA

1 10 100 1000 10 000concentration (ng g1)

Figure 8. Concentrations of organic contaminants in marineplastic debris (fragments). Solid diamond: The North Pacific

KongJap

San Francisco

Los Angeles

plastic pellets. Polychlorinated biphenyl concentration sum of138, 128, 187, 180, 170, 206.


on April 14, 2015http://rstb.royalsocietypublishing.org/Downloaded from correlated with those in mussels (R2 0.87). Thesedata clearly demonstrate that beached resin pellets canbe used to monitor pollution by hydrophobic chemicalson a global scale.

The hydrophobic surfaces of the resin pelletsaccumulate other chemicals in addition to PCBs. Sofar, 2,20-bis(p-chlorophenyl)-1,1,1-trichloroethane(DDT) and its metabolites (DDE and DDD), hexa-chlorinated hexanes (HCHs), PAHs and hopaneshave been detected in beached resin pellets.Interesting, regional distributions were observed.For example, higher concentrations of HCHs andDDTs were detected in South Africa and northernVietnam, respectively, probably because of the cur-rent use of these chemicals as pesticides in theseareas. These regional differences in contaminant

UK

Greece

Italy

Portugal

India

ThailandMalaysia

Indonesia

South AfricaMozambique

Australia

VietnamHong

The Netherlands

Figure 7. Concentrations of PCBs (ng g1 pellet) in beachedconcentrations of CB nos 66, 101, 110, 149, 118, 105, 153,concentrations in the resin pellets imply that ecologi-cal risks associated with the contaminants adsorbedto marine plastics will also vary among the areas.

Considering potential effects on the marine eco-system, plastic fragments, including microscopicfragments, are possibly more important than resin pelletsbecause fragments are more abundant among marineplastic debris (e.g. Moore et al. 2001; McDermid &McMullen 2004; see also Barnes et al. 2009; Ryanet al. 2009). Similar to resin pellets, hydrophobiccontaminants sorb to marine plastic fragments.Figure 8 shows the concentrations of various pollu-tants detected in plastic fragments collected from abeach near Tokyo, Japan and from the North CentralPacific Gyre (approx.1000 km off the west coast ofthe USA). The latter were floating plastic fragmentsand collected by a neuston net. The plastic fragmentswere sorted, and polyethylene fragments with variousshapes and sizes ranging from 1 10 20 to 31 35 35 mm were extracted with dichloromethane bySoxhlet for analysis. To evaluate variability, severalpools were analysed for each beach. As observedwith the resin pellets, PCBs, DDE and PAHs weredetected in plastic fragments from both areas.

Phil. Trans. R. Soc. B (2009)an

BostonConcentrations of contaminants in polyethylene frag-ments were higher on the Japanese coast (PCBs:12254 ng g1; DDE: 0.2276 ng g1; PAHs:,609370 ng g1) than those in the Central Gyre(PCBs: 123 ng g1; DDE: 0.14.7 ng g1; PAHs:,100959 ng g1). This difference can be explainedby regional differences in seawater concentrations ofthe contaminants (i.e. coast versus open ocean),since these compounds enter the aquatic environment

Central Gyre; open circle: Japanese coast of the PacificOcean. Polychlorinated biphenyls: sum of concentrations of

CB nos 66, 101, 110, 149, 118, 105, 153, 138, 128, 187,180, 170, 206; DDE: concentration of p, p0-dichlorodiphenyldichloroethene; PAHs: sum of concentrations of phenan-threne, anthracene, methylphenanthrenes (substitutionposition: 3, 2, 9, 1), fluoranthene, pyrene, benz[a]anthracene,

chrysene, benzo[b]fluoranthene, benzo[j]fluoranthene,benzo[k]fluoranthene, benzo[e]pyrene, benzo[a]pyrene,indeno[1,2,3-cd]pyrene, benzo[ghi]perylene and coronene;PBDEs: concentration of BDE nos 3, 7, 15, 17, 28, 71,49, 47, 66, 77, 100, 119, 99, 85, 126, 154, 153, 138, 183;

NP: concentration of nonylphenols; OP: concentration ofoctylphenol; BPA: concentration of bisphenol A.

bromines) were 0.92.1 ng g1 for the Japanese coastaddition, a number of platinum group metals have


on April 14, 2015http://rstb.royalsocietypublishing.org/Downloaded from and 0.457 ng g1 for Central Gyre. Since PBDEsare ubiquitous in coastal seawater and are hydrophobic,those on the plastic fragments may have sorbed fromsurrounding seawater. Alternatively, they may havebeen added to the plastics. In the plastic fragmentsamples from the Central Gyre, BD no. 183 was predo-minant. Since BD no. 183 is a major component ofOcta-BDE, which is a flame retardant currently usedpredominantly from terrestrial runoff. Rios et al.(2007) reported similar concentrations of PCBs,DDE and PAHs in marine plastics (both resin pelletsand fragments) collected from a wide variety ofPacific Ocean locations including the North PacificGyre, California and Hawaii. They also detectedDDT, DDD, HCHs and n-alkanes.

(b) Additive-derived chemicalsPlastics additives have also been measured on marineplastic debris. Nonylphenol, which forms from TNPdegradation (2), and is also present as an impurityin TNP (Gilbert et al. 1986), was detected in plasticresin pellets collected from 12 Japanese coasts and7 Malaysian coasts (Mato et al. 2002). Nonylphenolconcentrations ranged from 18 to 17 000 ng g1.Interestingly, no difference in NP concentrations wasobserved between Japanese and Malaysian coasts,although PCB concentrations in the pellets fromJapanese coasts were much higher than those from theMalaysian coast, probably because of the difference inpollution levels of PCBs in the coastal waters.Polypropylene pellets tended to have higher amountsof NP than polyethylene pellets (Mato et al. 2002),which is consistent with a higher prevalence of additivesin polypropylene pellets.

Since additives are also compounded into plasticsduring moulding, fragments of post-consumerplastics may have more additive-derived chemicalsthan resin pellets. Nonylphenol, OP and BPA weredetected in marine plastic fragments, as shown infigure 8. Interestingly, higher concentrations of NP(24.92660 mg g1) were found in the polyethylenefragments from the open ocean (Central Gyre) thanthe Japanese coast, where higher concentrations ofhydrophobic contaminants such as PCBs and PAHswere found. It is therefore likely that these NPs arederived from additives rather than adsorption fromsurrounding seawater. Significant concentrations ofBPA (5284 ng g1) were also detected in the plasticsamples from the Central Gyre, whereas no significantBPA was detected in plastic fragments from theJapanese coast (,25 ng g1). Because only poly-ethylene plastic was sorted and analysed (i.e. nopolycarbonate and epoxy resin was included), BPAdetected in the marine plastic fragments is also likelyto have been derived from an additive. Another impor-tant category of additive includes flame retardantssuch as polybrominated diphenyl ethers (PBDEs).Significant concentrations of PBDEs consisting ofmonobrominated to heptabrominated diphenyl etherswere detected in the plastic fragments both from theJapanese coast and the Central Gyre. Concentrationsof PBDEs (sum of BDEs ranging from three to sevenPhil. Trans. R. Soc. B (2009)demonstrated a high affinity for plastics (Cobelo-Garcia et al. 2007). This suggests a high potential forplastics to transport such chemicals. The nature andenvironmental significance of this transport are thenpartly determined by the subsequent fate of the plastic.Sorption of contaminants to plastics may also inhibitcontaminant biodegradation. For example, microbialdegradation of phenanthrene has been shown to bereduced by a factor of 6 when associated within USA and not predominant in US coastal waters,this congener in the plastic fragment from the CentralGyre most probably originated from the flame retardantadded to plastics. Polybrominated diphenyl ethers aresuspected thyroid disruptors in wildlife and humans(WHO/IPCS 1994). Therefore, more research isneeded to identify the sources of PBDEs in the plasticfragments.

6. TRANSFER OF PLASTIC-DERIVEDCONTAMINANTS TO ORGANISMSUptake of contaminants by organisms occurs by a var-iety of pathways, most commonly inhalation, dermalsorption and ingestion. Contaminant transfer to organ-isms from plastics may occur by any of these routes, andthe major transport route will vary according to theorganism and the physico-chemical properties ofthe contaminant. For most species, the predominantroute of transfer of contaminants from plastics is likelyto occur via plastic ingestion.

More than 180 species of animals have been docu-mented to ingest plastic debris, including birds, fish,turtles and marine mammals (Laist 1997). Smallplastic pieces floating on the ocean surface are mista-ken for food by both fish and birds, while turtles eatsuspended plastic bags, which they may mistake forjellyfish. To some extent, plastic ingestion by marinemammals may also be indirect, occurring by ingestionof fish that have eaten plastic (Eriksson & Burton2003). In addition, plankton trawls recover substantialquantities of plastic (Moore et al. 2001), and plankton-feeding mammals such as mysticetes (baleen whales)are also at risk of plastic ingestion (see Gregory2009). Temporal studies suggest that plastic ingestedby birds has been increasing since the 1960s(Robards et al. 1995; Vlietstra & Parga 2002; Ryanet al. 2009). Since few of these ingestion studies datelater than the early 1990s, it is difficult to assess theeffect of the enactment of later legislation such asMARPOL Annex V, which came into effect in 1989and is intended to restrict further the disposal ofplastics at sea (Derraik 2002). However, Cadee (2002)reported that 80 per cent of plastic debris on theDutch coast had peck marks made by birds, indicatingwidespread sampling of plastic by birds. Ingestion ofplastics has many detrimental consequences, includinggastrointestinal blockages (Baird & Hooker 2000),ulceration (Fry et al. 1987), internal perforation anddeath (Mascarenhas et al. 2004).

A lesser-studied problem is the transfer of contami-nants from ingested plastics to organisms. Sections 2and 5 detail the presence of a selection of organicand organometallic contaminants in plastics. In

w/w) for 10 days. At the end of the trial, contaminant


on April 14, 2015http://rstb.royalsocietypublishing.org/Downloaded from concentrations in the lugworm tissue were significantlyhigher than that in the sediment. While this demon-strated that contaminants can be transferred fromplastics to benthic organisms (R. C. Thompson 20042007, unpublished data), the mechanism remainsunclear. Some plastic debris is comparable in size tosediment and suspended particulate matter and isingested by invertebrates with varying feeding patterns.Sediment-ingesting lugworms, filter-feeding barnacles(Semibalanus balanoides) and amphipods (Orchestiagammarellus), which eat decaying organic matter, haveall been shown to ingest plastic fragments (Thompsonet al. 2004). Therefore, ingestion of contaminant-sorbed plastics may be a route of contaminant transportto these organisms. Indeed, digestive surfactants isolatedfrom deposit feeders have been shown to enhancedesorption of pollutants including PCBs and PAHs(Voparil & Mayer 2000; Ahrens et al. 2001).

Perhaps more important is the potential for invert-ebrates to uptake contaminants passively, as describedby equilibrium partitioning. Contaminants accumu-late in the tissues of some organisms as a consequencepolyethylene (Hatzinger & Alexander 1997). Thus,plastics not only have the potential to transportcontaminants, but they may also increase theirenvironmental persistence. Ingestion of plastics withsorbed contaminants has been suggested as a possibleexposure route for contaminants (Fry et al. 1987; Ryanet al. 1988; Mato et al. 2001; Thompson et al. 2004).In support of this hypothesis, the quantity of the con-taminant desorbed from the plastic was greatlyenhanced by the presence of surfactants and organicmatter (Sakai et al. 2000), suggesting that increasedleaching of contaminants will occur under gastricconditions. Furthermore, acidic gastric conditionsmay enhance desorption of metals bound to plastics.Finally, a positive correlation has been observedbetween the mass of ingested plastic and the PCB con-centration in the fat tissue of birds (great shearwaters;Puffinus gravis) (Ryan et al. 1988). Since plastics areknown to accumulate PCBs in the environment (5),this correlation supports plastic-mediated transfer ofcontaminants to organisms.

7. TRANSFER OF CONTAMINANTS FROMPLASTICS TO LOWER TROPHIC ORGANISMSMicroscopic plastic debris as small as 20 mm(Thompson et al. 2004) also litters the global environ-ment (Thompson et al. 2004; Ng & Obbard 2006;Reddy et al. 2006). This material enters the environ-ment both directly from use as scrubbers in cleaningproducts and as abrasive beads for cleaning ships, andindirectly from deterioration of brittle, weatheredmacroscopic plastics (Derraik 2002). To investigatethe possibility for microscopic plastics to transport con-taminants to benthic organisms, lugworms (Arenicolamarina) were exposed to contaminant-sorbed plastics.Phenanthrene (a PAH), tetrabromodiphenyl ether(a PBDE), triclosan (antimicrobial) and NP weresorbed to microscopic PVC particles at environmentallyrelevant concentrations. Lugworms were exposed tosediment containing these contaminated plastics (5%Phil. Trans. R. Soc. B (2009)of the organism reaching equilibrium with theirsurrounding environment. Determination of the equi-librium concentration of a contaminant between theanimal lipids and their environment (e.g. the sur-rounding water, sediment, soil) is an effectivemethod for estimating contaminant burden in biota(Di Toro et al. 1991; European Chemicals Bureau2003; Reichenberg & Mayer 2006). This method hasbeen widely applied to explain accumulation of bothorganic and inorganic contaminants in sediment andsoil-dwelling organisms (European Chemicals Bureau2003). Uptake probably occurs through both the skinand the gut, but ultimately the organism reachesequilibrium with its environment.

Using the equilibrium partitioning model, the effectof contaminant uptake from plastic into the lugwormwas examined (Teuten et al. 2007). Distribution co-efficients (Kd) were measured in vitro for sorption ofphenanthrene to a selection of sediments and plasticsincluding polyethylene, polypropylene and PVC fromseawater. Unsurprisingly perhaps, the measured Kdvalues were up to three orders of magnitude higherfor sorption to plastic than to sediment, with uptakeonto polyethylene exceeding that onto the otherplastics by an order of magnitude (Teuten et al.2007). These measurements were then used to esti-mate the amount of phenanthrene expected toaccumulate in lugworm tissue in the presence of plas-tic. Addition of clean polyethylene to sandy sediment,at plastic concentrations reported in the environment,was predicted to decrease the amount of phenanthrenein lugworm tissue by 13 per cent, compared withaccumulation in the absence of plastic (Teuten et al.2007). This can be viewed as scavenging of the phe-nanthrene by the highly sorbent plastic, which reducesthe equilibrium concentrations in the other phases,including the organism. However, owing to the highuptake of contaminants onto plastics and the longevityof plastics in the environment, it is unlikely that plasticdebris will remain clean for any extended period oftime. Hence any predicted beneficial effect of plasticsin reducing contaminant concentrations in benthicorganisms will likely be short-lived, if it operates atall (Teuten et al. 2007). Many plastics float in thesea-surface microlayer (SML), where contaminantconcentrations can be highly enriched over concen-trations in the bulk water column (Wurl & Obbard2004). Further fouling of floating plastics causesthem to become negatively buoyant (Ye & Andrady1991), which may result in sedimentation, and thusact as a transporter of contaminants into offshore aswell as strandline sediments.

Equilibrium partitioning was also used to examinethe effect of scavenging of phenanthrene in the SML,followed by deposition of the contaminated plasticsinto sediments, on subsequent phenanthrene accumu-lation by lugworms living in the sediment. Figure 9displays a schematic in which this transfer mechanismis compared with direct addition of clean plastic to thesediment, as described earlier. The calculationsshowed that only a small quantity of plastic would berequired to sorb a significant amount of phenanthrenefrom the SML and transport it into the sediment. Thiswould increase total phenanthrene in the benthic

((

s incum

seqentcur


on April 14, 2015http://rstb.royalsocietypublishing.org/Downloaded from environment and was predicted to result in higherlugworm tissue concentrations of phenanthrene.Compared with plastic-free sediment, addition ofonly 1 ppm of phenanthrene-contaminated polyethy-lene to sandy sediment was predicted to increase thephenanthrene tissue concentration of the worms by80 per cent. This, however, has yet to be verified byin vivo experiment.

Little is known about the effect of modification ofplastics by weathering on the sorption of contami-

air(a)

(b)

water

sediment

air

water

sediment

plastic particle

Figure 9. Schematic illustrating the additional effects of plasticto clean plastic in sediment resulting in (b) subsequent ac(c) sorption of phenanthrene to plastic in the SML and subin the sediment. Note that (c)!(d) results in higher sedimshown in the schematic, sorption to the sediment also ocCopyright 2007 (American Chemical Society).nants to plastics, but photo-oxidation is likely to beimportant as photolytic reactions are known tomodify plastic surfaces (Neidlinger & Schissel 1987).For example, polymer chain scission ultimately leadsto cracking, thus increasing the surface area of theplastics (Neidlinger & Schissel 1987). This may thenincrease uptake of contaminants. Reaction withoxygen increases the surface polarity of some plastics(Neidlinger & Schissel 1987), decreasing affinity forhydrophobic contaminants. The results of a series ofexperiments, complementing the above work anddesigned to investigate the effect of photo-oxidationon sorption of contaminants to polymers, areillustrated in figure 9.

Granules (150200 mm) of three high-productionplastics (polyethylene, polypropylene and PVC) wereartificially weathered using a filtered xenon lamp(Heraeus Suntest CPS), while maintaining the temp-erature at 25+38C in a glycerol bath connected to aCoolflow CFT-33 refrigerated recirculator. The spec-trum of this lamp is similar to that of natural sunlight(Ali et al. 1995). The plastics were agitated daily tomaximize exposure to radiation from the lamp.Calibration of the lamp by optometry indicated that1.08 h of exposure time was equivalent to 1 day ofFlorida sunlight (308N, 12 h) (West 2007). Plasticswere exposed to the lamp for 7 and 16 days. This

Phil. Trans. R. Soc. B (2009)was determined to be equivalent to 208 and 460days of exposure to sunlight at 308N. The plasticsdid not appear visibly altered by exposure to thelamp. Aged polypropylene showed an additionalabsorption at approximately 1750 cm21 as determinedby Fourier transform infrared spectroscopy (FTIR),characteristic of the presence of a carbonyl group(C = O), presumably because of oxidation. A similarabsorption was faintly visible in the FTIR spectrumof polyethylene aged for 16 days. Distribution co-

c)

d)air

water

phenanthrene

sediment

air

water

sediment

the transport of phenanthrene: (a) sorption of phenanthreneulation of phenanthrene in the sediment, compared with

uent sinking resulting in (d) accumulation of phenanthrenephenanthrene concentrations than (a)!(b). Although not

s (reproduced with permission from Teuten et al. 2007).efficients (Kd) were determined for sorption ofphenanthrene to these artificially weathered polymers,and tissue concentrations in lugworms were estimatedusing equilibrium partitioning, as described previously(Teuten et al. 2007).

Figure 10 shows the measured Kd for sorption ofphenanthrene to plastics from seawater and the pre-dicted amount of contaminated plastic required insediment with 0.2 per cent organic carbon to give an80 per cent increase in lugworm tissue, comparedwith plastic-free sediment. For polyethylene, a clearand statistically significant increase in Kd with lengthof exposure to light was observed (figure 10). Theresults for the other plastics were quite variable.However, in all cases Kd is lower for sorption tovirgin plastic than to plastic exposed to the xenonlamp for 16 days, so that less of the weathered plasticwould be required to transport phenanthrene to lug-worms. The most likely explanation for the variationis that two competing surface changes occur onphoto-oxidation, giving opposite effects. An increasein surface area will increase Kd, while an increase insurface oxidation might decrease Kd for hydrophobicorganic compounds. The increase in oxidation isparticularly evident for exposed polypropylene, forwhich FTIR analysis indicated the appearance of acarbonyl group arising from reaction of oxygen with

PVC


on April 14, 2015http://rstb.royalsocietypublishing.org/Downloaded from (b)PE

PP

PVC

0

0 10 000 20 000 30 000 40 000 50 000

5 10 15 20 25amount of plastic added to sediment (ppm)

duration under UV lamp 16 days 7 days 0 days

distribution coefficient (l kg1)

Figure 10. (a) Distribution coefficients (Kd) for sorption ofphenanthrene to UV-treated plastics from seawater.(b) Amount of phenanthrene-sorbed plastic required in sedi-ment (0.2% organic carbon) to increase lugworm tissueconcentration by 80%, compared with plastic-free sediment,predicted using equilibrium partitioning as describedpreviously (Teuten et al. 2007). Plastics were exposed to aUV lamp for 9 and 16 days, equivalent to 208 and 460days in natural sunlight. Note that plastic concentrationsPE(a)

PPthe polymer. This correlated with a decrease in Kdafter 7 days light exposure, indicating a reduction inthe amount of phenanthrene sorbed to polypropylene.Accurate determination of the surface area of the ultra-violet (UV)-exposed polymers was not possible withthe amounts used. Further work involving light treat-ment would be advantageous. These variable resultsare consistent with work by Endo et al. (2005) whoobserved no significant relationship between PCB con-centrations in beached plastic resin pellets and theircarbonyl index.

The discussion mentioned earlier illustrates thattransfer of contaminants from plastics to biota is a com-plex phenomenon relying on a variety of processes. Formany lower-trophic-level organisms, the extent of con-taminant uptake from plastics will be determined byequilibrium partitioning with the surrounding environ-ment. This may be facilitated by active desorption ofthe contaminants in vivo, for example, by solubilizationin digestive fluids. The nature and the history of theplastic will also be important in governing transfer ofcontaminants to organisms.

8. TRANSFER OF CONTAMINANTS TOHIGHER-TROPHIC-LEVEL ORGANISMS ANDBIOMAGNIFICATIONIn the previous sections, we have demonstrated thatchemicals adsorbed to the plastics may be released todigestive fluid and can be transferred to the tissue of

in sediment are well below the maximum reported amountof 81 ppm (Reddy et al. 2006).

Phil. Trans. R. Soc. B (2009)organisms. However, the situation for higher-trophic-level organisms is more complex because of biomagni-fication. Tissue concentrations of hydrophobic andpoorly metabolizable contaminants, such as PCBs,are amplified through the food web. Higher-trophic-level organisms (e.g. seabirds) are exposed to highlyenriched concentrations of hydrophobic contaminantsvia their prey (e.g. fish). Therefore, ingested plastics(i.e. anthropogenic prey) compete with the naturalprey in terms of contaminant burden to the predator.To assess the potential hazard of ingested plastic-derived chemicals to the predator, chemicals presentin the natural prey should be considered. This sectionreports the preliminary results of a field experiment inwhich seabirds were fed with plastic resin pellets toexamine the transfer of contaminants from the plasticto the seabird.

Chicks of streaked shearwater (Calonectris leucome-las) were fed with natural fish and kept in cages locatedat Mikurajima Island, a natural breeding ground of thebird. Eight 40-day-old chicks were used for the exper-iment. Among them, five individuals were fed withpolyethylene resin pellets collected from Kasai seasidepark in Tokyo Bay that contained significant amountsof PCBs as described later. Forty pellets (approx. 1 g)were mixed with the natural prey ( Japanese sandlance: Ammodytes personatus) at the beginning of theexperiment (day 0) and fed to each of the five chicks.Some of the pellets were not fed to the birds andwere stored for PCB determination. As a control,three individuals were fed with the Japanese sandlance without plastic resin pellets. For both plastic-feeding and control settings, approximately 10120 gwet of the fish was fed to each chick daily. Some ofthe fish were stored in a freezer until the PCBmeasurement. The experiment continued for 42 daysand every 7 days preen gland oil, which is secretedfrom the preen gland located at the base of the tailfeathers, was collected from the live chicks.Polychlorinated biphenyls in the resin pellets, the fishand the preen gland oil were analysed through solvent-extraction, chromatographic purification and a gaschromatograph equipped with an electron capture detec-tor (GC-ECD). Twenty-four PCB congeners (IUPACnumbers 8/5, 18, 28, 52, 44, 66/95, 90/101, 110/77,118, 132/153, 105, 138/160, 187, 128, 180, 170/190and 206) were identified and quantified. The total con-centration of these congeners is expressed as total PCBin the present study. The details of the analyticalprocedure for the plastic pellets and fish were describedin Endo et al. (2005) and for preen gland oil inYamashita et al. (2007).

In the pellets from Tokyo Bay, PCBs were detectedat concentrations ranging from 51 to 562 ng g1, witha mean of 97 ng g1. Based on the number of pelletsfed to the streaked shearwaters, each chick is estimatedto have been exposed to approximately 100 ng ofPCBs. Significant concentrations of PCBs were alsodetected in the fish fed to the chicks, because thesefish bioconcentrate PCBs through their prey (e.g.copepods). Polychlorinated biphenyl concentrationsin the fish ranged from 0.298 to 0.706 ng g1 wet(mean: 0.451 ng g1 wet). The PCBs in the fishshould be compared with those in the pellets to

0.5

021

rela

tive

c

PCBs

in

elapsed time (d)0 7 14 28 35 42

Figure 12. Time-course of PCBs in preen gland oil of thechicks during the feeding experiment: (a) total PCBs* and(b) lower chlorinated congeners**. *Total PCBs: sum ofCB nos 8, 5, 28, 52, 44, 90, 101, 110, 77, 118, 132, 153,

138, 160, 187, 128, 180, 170, 190, and 206; **lower chlori-nated congeners: sum of CB nos 8, 5, 18, 28, 52, 44, 66, 95.Polychlorinated biphenyl concentrations are normalized tothose on day 0 on each series. Closed symbols and solid

lines: plastic-feeding setting; open symbols and dottedlines: control setting. Replicate chicks no. 1 (closedsquare), 5 (cross), 8 (closed diamond), 10 (closed circle),14 (closed triangle), 4 (open circle), 9 (open triangle), 18(open diamond) were analysed. For chicks 1, 5, 8, 4, 18,

samples were analysed on day 0 and day 7 only.


on April 14, 2015http://rstb.royalsocietypublishing.org/Downloaded from evaluate the contribution of the plastics to the totalPCB burden of the seabird. Based on the weight ofdaily feeding of fish (approx. 10120 g), dailyexposure of PCBs from the fish is calculated to beapproximately 15 ng. This daily exposure (approx.15 ng) is lower than those from the resin pellets(approx. 100 ng). However, plastic was fed once (day0) during the experiment, while fish were fed dailythroughout the experiment. Therefore, in the earlystage of the experiment (days 07), PCB exposurefrom the pellets dominated those from the fish,whereas after Day 7, PCB exposure from the fish over-came that from the plastics (figure 11). This suggeststhat potential effects of fed-pellet-derived PCBscould be observed in the early stage of the experiment.

Figure 12 shows the preliminary results on the time-course of PCB concentrations in the preen gland oilfrom the streaked shearwater. It has been previouslydemonstrated that PCBs are accumulated in the preengland oil owing to their hydrophobic nature and thatthe PCB concentrations in the oil reflect PCB levels inthe internal tissue (i.e. abdominal adipose; Yamashita

00

200

400

600

800

1000

1200

1400

7 14 21 28 35 41(d)

cum

ula

tive

load

of P

CB (n

g)

Figure 11. Loads of total PCBs in chicks. Closed circles:

cumulative load from fish ( Japanese sand lance: Ammodytespersonatus); dotted line: load from plastic resin pellets.et al. 2007). Preen gland oil can be taken from a livebird without injury or stress. It provides a practical anduseful tool to monitor PCB concentrations in biologicaltissue of live seabirds over time. Figure 12 displays therelative concentrations of PCBs in the individualchicks. For each chick, the preen gland oil PCB concen-tration is normalized to that at day 0. The PCB concen-trations in preen gland oil increased through day 7 for theplastic-feeding setting, whereas no such increase wasobserved in the control. On the other hand, in the laterstage, PCB concentrations in preen gland oil showedno difference between the plastic-feeding and the controlsettings (figure 12a). This is consistent with a larger PCBdose from the plastics than from fish-only in the earlystage of the experiment (through day 7). These resultssuggest that plastic-derived PCBs were transferred tothe biological tissue of the seabird. However, the differ-ence between the plastic-feeding chicks and the controlwas not statistically significant.

To obtain clearer evidence on the nature of thetransfer, we focused on lower chlorinated congeners,

Phil. Trans. R. Soc. B (2009)0

3.0

2.5

2.5

2.0

2.0

1.5

1.5

1.0

1.0

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07 14 21 28 35 42

elapsed time (d)

(b)

(a)

rela

tive

conce

ntr

atio

n of

PCB

in p

reen

gla

nd o

ilonce

ntr

atio

n of

pre

en g

land

oilbecause lower chlorinated congeners were relativelyrich in plastic resin pellets compared with the fish(figure SI in the electronic supplementary material).The difference was probably because of biomagnifica-tion of higher chlorinated congeners in fish throughthe food web. Lower chlorinated congeners are moreeasily metabolized and selectively depleted whenPCBs move through the food chain and, consequently,are less abundant in the fish tissue. Plastic resin pellets,however, just concentrate (partition) PCB congenersfrom seawater (not through the food web) and nometabolic process occurs, therefore lower chlorinatedcongeners that could be subject to biological degra-dation are not depleted. Lower chlorinated congenerscan be regarded as a sensitive tracer to detect the con-tribution from plastic-derived PCBs. As illustrated infigure 12b, concentrations of lower chlorinated conge-ners in the plastic-feeding chicks increased up to threetimes from day 0 to day 7, whereas no increase in thecongeners was observed for the control. This differ-ence was statistically significant (p, 0.05, two-tailed

In higher-trophic-level organisms, plastic-derivedcontaminants compete with the biomagnified contami-

uptake of contaminants than other plastics includingpolypropylene and PVC. Ultraviolet weathering also


on April 14, 2015http://rstb.royalsocietypublishing.org/Downloaded from appears to affect the uptake of contaminants by plas-tics. Unfortunately, our understanding of how toevaluate the risks associated with chemicals derivedfrom marine plastics is poor. However, evidence isemerging that plastics with environmental contami-nants can transport these compounds to organismsat various trophic levels.

Finally, these findings warn us that we should notunderestimate the environmental impact of discardedplastics. As plastics production and usage continueto increase, particularly in economically developingcountries, the environmental implications of theirdisposal should be carefully considered to avoidinadvertent release, magnification and transport ofcontaminants (Thompson et al. 2009a,b).

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