controlling persistent organic pollutants–what next?

Environmental Toxicology and Pharmacology 6 (1998) 143–175

Assessment

Controlling persistent organic pollutants–what next?1

Harry W. Vallack a, Dick J. Bakker b, Ingvar Brandt c, Eva Brostrom-Lunden d,Abraham Brouwer e, Keith R. Bull f, Clair Gough g, Ramon Guardans h, Ivan Holoubek i,Bo Jansson j, Rainer Koch k, Johan Kuylenstierna a, Andre Lecloux l, Donald Mackay m,

Patrick McCutcheon n, Paolo Mocarelli o, Rob D.F. Taalman p

a Stockholm En6ironment Institute at York, Uni6ersity of York, York, United Kingdomb TNO Institute of En6ironmental Sciences, Energy Research and Process Inno6ation, Den Helder, The Netherlands

c Department of En6ironmental Toxicology, Uni6ersity of Uppsala, Uppsala, Swedend Swedish En6ironmental Research Institute, Goteborg, Sweden

e Department of Food Technology and Nutritional Sciences, Di6ision of Toxicology, Agricultural Un6ersity, Wageningen, The Netherlandsf Institute of Terrestrial Ecology, Huntingdon, United Kingdom

g Institute for Systems Engineering and Informatics, Joint Research Centre, Ispra, Italyh Centro de In6estigacione Energeticas MedioAmbientales e tecnologecas (CIEMAT), Madrid, Spain

I Department of En6ironmental Chemistry and Ecotoxicology, Masaryk Uni6ersity, Brno, Czech Republicj Institute of Applied En6ironmental Research, Stockholm Uni6ersity, Stockholm, Sweden

k Product Safety Department, Bayer AG, Le6erkusen, Germanyl European Chemical Industry Council (CEFIC) – EURO CHLOR, Brussels, Belgium

m En6ironmental Modelling Centre, Trent Uni6ersity, Ontario, Canadan European Commission, DG XI, En6ironment, Brussels, Belgium

o Facolta di Medicina e Chirurgia, Uni6ersita degli Studi di Milano, Desio-Milano, Italyp European Chemical Industry Council (CEFIC)–Endocrine Modulators Steering Group (EMSG), Brussels, Belgium

Accepted 7 September 1998

Abstract

Within the context of current international initiatives on the control of persistent organic pollutants (POPs), an overview isgiven of the scientific knowledge relating to POP sources, emissions, transport, fate and effects. At the regional scale,improvements in mass balance models for well-characterised POPs are resulting in an ability to estimate their environmentalconcentrations with sufficient accuracy to be of help for some regulatory purposes. The relevance of the parameters used to definePOPs within these international initiatives is considered with an emphasis on mechanisms for adding new substances to the initiallists. A tiered approach is proposed for screening the large number of untested chemical substances according to their long-rangetransport potential, persistence and bioaccumulative potential prior to more detailed risk assessments. The importance of testingcandidate POPs for chronic toxicity (i.e. for immunotoxicity, endocrine disruption and carcinogenicity) is emphasised as is a needfor the further development of relevant SAR (structure activity relationship) models and in vitro and in vivo tests for these effects.Where there is a high level of uncertainty at the risk assessment stage, decision-makers may have to rely on expert judgement andweight-of-evidence, taking into account the precautionary principle and the views of relevant stake-holders. Close co-operationbetween the various international initiatives on POPs will be required to ensure that assessment criteria and procedures are ascompatible as possible. © 1998 Elsevier Science B.V. All rights reserved.

Keywords: Persistent organic pollutants; Bioaccumulation; Biomagnification; Persistence; Global distillation; Risk assessment;Monitoring; Chronic toxicity; Endocrine disruption; Immunotoxicity; Carcinogenicity

1 Correspondence to: Nicole G Weavers, ETAF, European Training and Assessment Foundation, PO Box 182, 6700 AD, Wageningen, TheNetherlands. Tel.: +31-317-484924; Fax: +31-317-484941

1382-6689/98/$ - see front matter © 1998 Elsevier Science B.V. All rights reserved.PII S1382-6689(98)00036-2

H.W. Vallack et al. / En6ironmental Toxicology and Pharmacology 6 (1998) 143–175144

Table 1Priority substances in the CLRTAP POPs protocol (substances also specified in UNEP Governing Council Decision 19/13c for initial inclusionin the global POPs convention are shown in bold type)

Pesticides Industrial products Unintentional by-products of combustion and industrial processes

Polycyclic aromatic hydrocarbons (PAHs)cHexabromobiphenylaAldrina

Chlordanea Polychlorinated biphenyls (PCBs)a,b Polychlorinated dibenzo-p-dioxins (PCDDs)c

Chlordeconea Polychlorinated dibenzofurans (PCDFs)c

DDTa,b

Dieldrina

Endrina

Heptachlora

Hexachlorobenzene (HCB)a,c

Hexachlorocyclohexane(HCH)a,b

Mirexa

Toxaphenea

a Listed under Annex I of the protocol (substances scheduled for elimination).b Listed under Annex II of the protocol (substances scheduled for restrictions on use).c Listed under Annex III of the protocol (substances scheduled for emission reductions by the use of BAT (best available technology) etc.).

1. Introduction

1.1. Definition

Organic substances that are persistent, bioaccumula-tive and possess toxic characteristics likely to causeadverse human health or environmental effects arecalled PBTs (Persistent, Bioaccumulative, Toxic sub-stances). In this context, ‘substance’ means a singlechemical species, or a number of chemical species whichform a specific group by virtue of (a) having similarproperties and being emitted together into the environ-ment or (b) forming a mixture normally marketed as asingle product. Depending on their mobility in theenvironment, PBTs could be of local, regional or globalconcern. Under the auspices of the United NationsEconomic Commission for Europe (UN-ECE) Conven-tion on Long-Range Transboundary Air Pollution(CLRTAP), a protocol on persistent organic pollutants(POPs) has been drawn up in which POPs are definedas ‘a set of organic compounds that: (i) posses toxiccharacteristics; (ii) are persistent; (iii) are liable tobioaccumulate; (iv) are prone to long-range atmo-spheric transport and deposition; and (v) can result inadverse environmental and human health effects atlocations near and far from their sources’ (UN-ECE,1998a). In other words, POPs are a subclass of PBTsthat are prone to long-range atmospheric transport anddeposition. The global extent of POP pollution becameapparent with their detection in areas such as theArctic, where they have never been used or produced,at levels posing risks to both wildlife (Barrie et al.,1992) and humans (Mulvad et al., 1996). Growingconcern over recent decades about the potential effectsof some man-made chemical substances on humanhealth and the environment has prompted action at

many levels from local to global. Some of the interna-tional initiatives that address these problems are de-scribed below in order to identify the main policy issuesand management tools that might contribute to moreeffective regulation and management of thesechemicals.

1.2. Regional initiati6es

1.2.1. The CLRTAP POPs protocolAn initiative on POPs was started within the UN-

ECE region (comprising eastern and western Europe,Canada and the United States) in 1992 with the estab-lishment of a Task Force on POPs within the frame-work of the 1979 Convention on Long RangeTransport of Air Pollution (CLRTAP); the Conventioncurrently having 43 parties. In 1996, the ExecutiveBody of the Convention established a PreparatoryWorking Group that drafted a comprehensive negotiat-ing text for a protocol and established a procedure foridentifying priority substances to be addressed by it.Starting from an initial list of 107 substances, 16 prior-ity substances were identified for initial inclusion in theprotocol comprising 11 pesticides, 2 industrial productsand 3 unintentional by-products (Table 1). Excludedfrom this list are compounds for which insufficient dataexisted and compounds for which there was no evi-dence of long-range atmospheric transport.

In 1997, parties to the Convention, meeting in itsnegotiating body, the Working Group of Strategies,started negotiations and in June1998, the final draftprotocol was presented to the Executive Body for adop-tion. The protocol is structured in 20 articles and eightannexes. The objective of it is ‘to control, reduce oreliminate discharges, emissions and losses of persistentorganic pollutants’. The obligations for the parties to itinclude:

H.W. Vallack et al. / En6ironmental Toxicology and Pharmacology 6 (1998) 143–175 145

1. to eliminate the production and use of the sub-stances listed in annex I in accordance with theimplementation requirements specified in it;

2. to ensure that the destruction or disposal of thesubstances listed in annex I is undertaken in anenvironmentally sound manner;

3. to restrict the substances listed in annex II to theuses described in it;

4. to reduce total annual emissions of the substanceslisted in annex III; and

5. to develop and maintain emission inventories for thesubstances listed in annex III.

The designations of the 16 substances in terms ofthese annexes are included in Table 1.

Some exemptions are allowed because legitimate usesexist for which there are no available alternatives atpresent (e.g. DDT for encephalitis and malaria control).These exemptions will be regularly reviewed in the lightof future technical developments.

Other articles of interest in this context are: article5—exchange of information and technology; article6—public awareness; article 8—research, developmentand monitoring and article 9 in which parties acceptobligations that will provide useful components forfuture development of knowledge about the risk ofPOPs and alternatives to their uses. Finally, article 14deals with amendments to the present protocol andincludes a reference to the detailed procedure, given ina separate document (EB decision 1998/2), developed tochange the list of priority substances (see Section 8, Fig.3).

1.2.2. Other regional programmes and initiati6esaddress the issue of POPs within the broader contextof hazardous substances

1.2.2.1. The Esbjerg declaration. In the Esbjerg Declara-tion of the 4th International Conference on the Protec-tion of the North Sea in June 1995, ministers agreed tothe objective of ‘continuously reducing discharges,emissions and losses of hazardous substances, therebymoving towards the target of their cessation within onegeneration (25 years) with the ultimate aim of concen-trations in the environment near background values fornaturally occurring substances and close to zero con-centrations for man-made synthetic substances’. As aninterim objective it confirmed the goal of ‘reducing bythe year 2000, discharges and emissions of substanceswhich are toxic, persistent and liable to bioaccumulate(especially organohalogen substances) and which couldreach the environment, to levels that are not harmful toman or nature with the aim of their elimination’.

Subsequently, the targets in this declaration havebeen adopted by both OSPAR (Oslo–Paris Conventionfor the Protection of the Marine Environment of theNorth-East Atlantic) and HELCOM (Helsinki Conven-

tion for the Protection of the Baltic) in their respectivestrategies on hazardous substances agreed in their 1998ministerial meetings. Similar targets have been adoptedby the Arctic Environment Protection Strategy of theCircumpolar Nations and the Barcelona Resolution,adopted by BARCOM (Barcelona Convention for theProtection of the Mediterranean Sea against Pollution).

1.2.2.2. The North American Commission for en6iron-mental co-operation (NACEC). Within the frameworkof the North American Free Trade Area (NAFTA)several North American Regional Action Plans(NARAPs) have been established for DDT, chlordaneand PCBs. The main objectives of the PCBs NARAPare to: (a) work toward the virtual elimination of PCBsin the environment, which the task force is interpretingas no measurable release to the environment, and thephase-out of uses for which release cannot be con-tained; and (b) propose environmentally sound man-agement and control of existing PCBs, throughout theirlife cycles, with special emphasis given to trans-boundary shipment of PCBs for disposal/destructionpurposes.

The main objective of the DDT NARAP is to reducethe exposure of humans and the environment to DDTand its metabolites through the phased reduction, andeventual elimination of DDT use for malaria controland the elimination of illegal uses of DDT. DDT iscurrently permitted for limited governmental use formalaria control in Mexico. The DDT NARAP willbuild on Mexico’s very successful malaria control pro-gram, and through an integrated pest-management ap-proach, continue to bring about reductions in the use ofDDT.

The objective of the Chlordane NARAP is to reducethe exposure of humans and the environment to chlor-dane through the phase-out of existing registered usesof chlordane. The chemical currently has limited use inthe control of termites. The NARAP will reflect anintegrated pest management approach, including themanagement of existing stocks and the phase-out ofchlordane use in North America.

1.2.3. ConclusionIt can be concluded that many parallel actions on

POPs are being initiated which provide a very strongdrive for action but they will also require urgent effortsto avoid duplication and enhance co-ordination acrosscountries, regions, institutions and social actors.

1.3. Global initiati6es on POPs

In the May 1995 United Nations Environment Pro-gram (UNEP) Governing Council (GC) decision 18/32,the Inter-organisation Programme for the Sound Man-agement of Chemicals (IOMC) was requested to initiate


an assessment process on POPs, starting with the 12substances shown (in bold) in Table 1. As a result ofthis, the International Forum for Chemical Safety(IFCS) concluded, at its meeting in Manila in June1996, that there was sufficient information to demon-strate that international action, including a global le-gally binding instrument, will be required to reduce therisks to human health and the environment arisingfrom the release of the 12 POPs. The IFCS recom-mended to the UNEP GC and the World Health As-sembly (WHA), that immediate international actionshould be initiated to protect human health and theenvironment from these 12 POPs. Their recommenda-tions were endorsed by the UNEP GC and the WHA in1997. At the request of the UNEP GC, an Intergovern-mental Negotiating Committee (INC) was establishedwith a mandate to prepare, preferably by the year 2000,an international legally binding instrument for imple-menting international action to reduce and/or eliminatethe emissions and discharges of these 12 POPs. Its firstsession was held in Montreal, Canada, 29 June–3 July1998.

The INC was requested to establish an expert groupfor the development of science-based criteria and aprocedure for identifying additional POPs as candidatesfor future international action. UNEP has also initiateda number of immediate actions involving developmentof alternatives to POPs; identification and inventoriesof PCBs; available destruction capacity; identificationof sources of PCDD/Fs and aspects of their manage-ment. Furthermore, UNEP has convened jointly withthe IFCS a series of awareness-raising workshops indifferent regions to help governments prepare for thesenegotiations and identify POPs-related issues at na-tional and regional levels.

In May 1997, the WHA also endorsed the recom-mendations of the IFCS and adopted a resolution onPOPs, which interalia calls on member states:� to involve appropriate health officials in national

efforts to follow up and implement decisions of theUNEP and WHO governing bodies relating toPOPs;

� to take steps to reduce reliance on insecticides forcontrol of vector borne diseases through promotionof integrated pest management approaches in accor-dance with WHO guidelines and through support forthe development and adoption of viable alternativemethods of disease vector control; and

� to ensure that DDT is authorised by governmentsfor public health purposes only and limited to gov-ernment-authorised programmes.The UNEP Governing Council decision 18/31 from

1995 concerns the Protection of the Marine Environ-ment from land-based activites through a global pro-gramme of action (GPA). The decision specificallyaddressed POPs and called on States to consider how,

within the GPA, appropiate attention could be given tosupport action at a national and regional level onPOPs. The GPA was adopted in November 1995 andincludes commitments to develop a global, legally bind-ing instrument for the reduction and/or elimination ofemissions, discharges and, where appropiate, the elimi-nation of the manufacture and use of the POPs iden-tified in decision 18/32.

1.4. Challenges in effecti6ely implementing pro6isions inthe CLRTAP protocol and the planned UNEP globalinitiati6e

The UN-ECE CLRTAP protocol is currently theonly international instrument for the control of risksfrom POPs as a distinct class of chemicals. It representsa comprehensive set of provisions for the elimination orrestriction in use of a number of POPs that can providea good basis for negotiations on the global agreementunder UNEP. In this section, challenges in effectivelyimplementing the provisions of the CLRTAP POPsprotocol are explored with reference to both the UNECE region and the global scale.

The CLRTAP protocol focuses on the atmosphericpathway as a transport route for POPs. Although thispathway is clearly one of the main routes by whichPOPs reach remote regions, other (in particularaquatic) pathways are likely to be considered for theglobal instrument. Many of the substances identified inthe protocol have already been restricted or phased outwithin UN-ECE countries. Consequently, implementa-tion of the obligations under the protocol will be lessproblematic for parties to the CLRTAP than for manycountries outside this region where these substancestend to be more widely used.

The CLRTAP POPs protocol identifies a number ofsubstances for elimination (annex I to the protocol).Some exemptions to these obligations are identifiedwhich allow production or use in limited circumstances.These are included to accommodate some countrieswhere significant current usage of some POPs precludesimmediate phase out. The challenge will be to ensurethat actual usage is limited to those authorised activitiesonly. Within the UN-ECE region there will be a needfor support and advice for countries in eastern Europeto help them to develop practical implementation plansand monitor compliance. Monitoring compliance forrestricted use will be much more difficult than forelimination. The success of monitoring efforts will de-pend upon access to trained personnel and institutionaldevelopment.

At the global scale, many countries use more of thePOPs listed in the CLRTAP protocol than do UN-ECEcountries and may have more difficulties with the elim-ination of these substances. Also, considerable chal-lenges may present themselves in the implementation of


restricted use of POPs in developing countries. One keyquestion is whether viable alternatives exist for all thedifferent situations in which POPs are current used indeveloping countries. Any phasing out of POPs willhave to go hand-in-hand with capacity building withrespect to infrastructure, legislative methods of controland monitoring. Particular problems can result fromold stockpiles of POPs, in particular DDT. The elimi-nation of the existing stocks in an environmentallysound manner will be a challenge in some parts of theworld. Illegal (non-essential) uses and illegal trade mayalso present major difficulties. There will be a need foradvice, information and training on implementing re-placements for current POPs usage. These may com-prise the use of alternative chemicals, the adoption ofenvironmental, other technical, or medical managementoptions for controlling pests and vector-borne diseasesas well as changes in methods of use, such as moreefficient application of DDT for malaria control. Dif-ferent control options for managing pests and vectorsare usually most effective when combined with, andintegrated into, strategies which are ‘customised’ toaddress specific situations. On-going activities in thisregard include the newly established Global IntegratedPest Management facility of the Food and AgricultureOrganisation (FAO)/UNEP/UN Development Pro-gram/World Bank and the WHO/FAO/UNEP Panel ofExperts on Environmental Management for VectorControl.

The UN-ECE CLRTAP obligations for PAHs,PCDD/Fs and HCB are phrased in terms of limitvalues for the potential sources, application of BAT tostationary and mobile sources and reductions in totalnational emissions. The challenge will be to encouragethe implementation of BAT where factors such as costsand effective transfer of technologies will be important.Structural changes in certain practices and lifestylesmay also lead to significant emission reductions. How-ever, there are risks that when old technology is re-placed it may be passed on to countries outside theUN-ECE area.

The need for emission inventories for POPs is iden-tified in the CLRTAP protocol. At present, few emis-sion inventories are available even for the UN-ECEregion and these are uncertain with some sources miss-ing. The choice of the base year identified under theprotocol (it can be between 1985–1995) may allow foremission increases from current levels for some coun-tries. In countries outside the UN-ECE region, emissionsources might be very different. For example, wasteincineration may not represent a major source in somecountries whereas in others, where open incineration iscommon, it would be much more significant. In general,facilities for the disposal of hazardous waste may notbe readily available in many of these countries. This

situation will not change without the introduction ofnew technologies and trained personnel. The capacityto develop emission inventories will have to be en-hanced both within the UN-ECE region and, evenmore so, on a global scale for the UNEP process.Information on relevant pathways, natural sources andsinks, is also missing.

Articles described in the CLRTAP protocol referringto exchange of information, public awarness and re-search and development, can contribute very signifi-cantly to the achievement of its objectives and thefuture development of fair and effective control optionsfor POPs. Consequently, it would be extremely cost-ef-fective in the medium and long term to promote theseactivities at both national and international levels. Thechallenge will be to put the ideas and issues identified inthese articles into practice. How exactly will the activi-ties be organised and implemented? Articles referring topublic awareness represent new ground for the CLR-TAP and it will be important to determine how theseinititiatives are to be organised and co-ordinated. Theimplementation of article 14 on amendments and theexecutive body decision on criteria and procedures toadd new substances to the protocol will be difficult.POPs are being considered under a number of interna-tional initiatives and there is a risk that scarce resourcescould be inefficiently used in parallel processes. There istherefore, a need for close co-operation between thedifferent international initiatives to ensure for example,that assessment criteria and procedures are as compat-ible as possible.

A major objective of this paper is to provide scientificguidance in the definition and use of assessment criteriaand procedures in order to ensure consistency through-out the various international and regional initiatives onPOPs.

2. Sources of POP emissions

2.1. Introduction

Emission inventories for POPs are restricted to rela-tively few compounds, are generally uncertain and lackspatial and temporal resolution (UN-ECE, 1994). Thesource of entry into the environment for manufacturedPOPs is determined by where and how they are used.Pesticides are intentionally released at their point ofapplication. Industrial chemical POPs are unintention-ally released by volatilization, leakage or leaching eitherduring a product’s lifetime or after ultimate disposal.Unintentionally produced POPs are by-products of in-dustrial or combustion processes emitted from bothstationary and mobile sources.


2.2. Pesticides

In many countries, there has been a gradual movetowards the use of less persistent and less bioaccumula-tive pesticides. However, in some developing countries,especially in tropical regions, organochlorine pesticides(such as DDT, HCH, chlordane and heptachlor) arestill used in agriculture and to control arthropods ofmedical or veterinary importance. This is reflected indeclining levels of HCH and DDT in human tissuesfrom developed countries whilst in many developingcountries these levels remain stable (Loganathan andKannan, 1994). The use of several organochlorine pesti-cides (particularly DDT) is still recommended by na-tional and international health organisations to controlmosquitoes, flies and lice which spread malaria, typhusand typhoid. However, the WHO predict that withreduced world-wide production, the rapid developmentof insect resistance and the possible appearance ofsuitable alternatives, DDT use may cease without sig-nificant legal intervention (WHO, 1996). Chlordane andheptachlor are still used to control termites althoughtheir use is steadily being phased out in one countryafter another.

Plants can either absorb these chemicals directlythrough their leaves or indirectly from the soil. Theplants may then be eaten by herbivores and accumulateto relatively high levels in meat and animal dairy prod-ucts. Pesticide residues in food create an important,though under-reported, exposure route (Repetto andBaliga, 1996).

Pesticides that are not bound in soils or taken up intoplants and animals can drain into rivers and lakes andmove into the aquatic food chain. Several chlorinatedpesticides have been detected in rivers in Tanzania,Colombia, Indonesia, Malaysia, China and Thailand atlevels suggesting potentially severe damage (Repettoand Baliga, 1996). The semi-volatile nature of thesepesticides also allows them to be dispersed rapidlythrough air. Iwata et al. (1994) showed that persistentand semi-volatile compounds (including HCHs andDDTs) discharged in the tropics tend to be redis-tributed on a global scale. In a study of the globaldistribution of persistent organochlorine pesticides, Si-monich and Hites (1995) reported high concentrationsnot only in some developing countries but also inindustrialised countries where the use of many of thesecompounds have been restricted for many years. Forexample, the presence of DDT, chlordane, dieldrin,HCHs and toxaphene in southern Ontario has beenlinked with their use in southern USA, Mexico and theCaribbean (Hoff et al., 1992). The latitudinal distribu-tion of atmospheric HCHs and DDTs suggests thattheir major contamination source in global terms hasshifted from mid to low latitudes during the 1980s(Iwata et al., 1993). Weber and Goerke (1996) also

found there had been a significant increase in p,p %-DDE(a metabolite of DDT) in various Antarctic fish speciesbetween 1987 and 1991. This is consistent with the factthat usage of many organochlorine pesticides is contin-uing, and in some cases increasing, in developing coun-tries near the tropics.

In addition to emissions from localised sources repre-sented by treated crops etc., revolatilization from largercontaminated areas, including oceans, soils and forests,also occurs. The cycling of these semi-volatile com-pounds between environmental compartments makes itdifficult to distinguish source from sink. Evaporationduring the warmth of the day can change to net deposi-tion with the cold of the night and seasonal tempera-ture changes can have a similar effect, albeit on anannual cycle (Hornbuckle and Eisenreich, 1996). Also,on a longer time scale, the decline in atmospheric HCHover the Arctic Ocean during recent years has reversedthe net direction of air–sea gas exchange to the pointwhere some northern waters are now sources of thepesticide to the atmosphere instead of sinks (Bidlemanet al., 1995).

2.3. Industrial products

2.3.1. Polychlorinated biphenylsThe polychlorinated biphenyls (PCBs) are a group of

chlorinated hydrocarbons consisting of 209 possiblecongeners ranging from three monochlorinated isomersto the fully chlorinated decachlorobiphenyl isomer.Their physico–chemical properties vary with respect totheir degree of chlorination. Generally, water solubility,vapour pressure and biodegradability decrease with anincreasing degree of chlorine substitution while hydro-phobicity and sorption tendency increase. Their combi-nation of high persistence and mobility means thatPCBs have been identified in almost every environmen-tal compartment or matrix (Fiedler, 1997). PCBs wereformerly used as dielectric fluids in transformers andlarge capacitors, as pesticide extenders, plasticisers insealants, as heat exchange fluids, hydraulic lubricants,cutting oils, flame retardants, dedusting agents, and inplastics, paints, adhesives and carbonless copy paper.Some of their applications resulted in direct or indirectreleases into the environment and large amounts werereleased due to inappropriate disposal, accidents (e.g.transformer fires) and leaks from industrial facilities.Although PCB production in most countries wasbanned in the 1970s and 1980s, the current worldinventory of PCBs is estimated at 1.2 million tonneswith about one third of this quantity circulating in theenvironment (Duursma and Carroll, 1996). Leakagefrom old equipment, building materials, stockpiles andlandfill sites constitutes a continued threat of PCBemission and some production is reported for certaincountries with economies in transition (UN-ECE,


Table 2Annual emissions of PAH, PCB, HCH and PCDD/Fs in Europe per source category in 1990 (adapted from Berdowski et al., 1997)

PCB (t/y)PAH (t/y) PCDDs/PCDFs (g I-TEQa)HCB (t/y)Source category

43040.46Fuel combustion (excluding transport) 6777 5.69Production processes :

1.46 0.24Iron and steel industry 612 1964— 0.05Non-ferrous metal industry 1010 1611

0.56— 0.2—Organic chemical industry— — — 4.5Paper and pulp industry

— 1.0Road paving with asphalt 112 —— —Cement industry 0.1 19.1— — 0.4—Glass industry— —Paint use 593 —

0.4——4817Wood preservation— 0.21Other solvent use — —0.08 —Road transport combustion 1488 73.1

63——195Other transport combustionWaste treatment and disposal :

0.07 2657Waste incineration 6 0.05— —Landfill — 19.9— — 23.1—Cremation— 5.76Pesticide use ——

——111—Electrical equipment— —Natural sources 2.153

a International toxic equivalents

1998a). Some time-trend surveys of PCB concentrationsin human adipose tissues show no significant declinethus implying continued exposure of humans, even indeveloped nations (Loganathan and Kannan, 1994).Conversely, decreasing trends have been observed intissues of various biota (e.g. fish, seals, birds) of theBaltic Sea (Olsson et al., 1997).

2.3.2. HexachlorobenzeneHexachlorobenzene (HCB), in addition to its use as a

fungicide, has been used in the manufacture of militarypyrotechnics, as a fluxing agent in aluminium smelting,as a porosity control in the manufacture of graphiteelectrodes, as a peptising agent in the rubber industryand as an intermediate in dye manufacture. It is also aby-product of the manufacture of various chlorine-con-taining chemicals and is a known impurity in severalpesticide formulations. HCB as a pesticide, is banned inmany countries and is severely restricted or voluntarilywithdrawn in several others. Estimates of the currentscale of international production are contradictory.HCB is ubiquitous in the environment and has beenmeasured in foods of all types and has been detected inArctic air, water and organisms (Ritter et al., 1995).

2.4. By-products

Emission data for developing countries and countriesin transition are scarce because of the costly analysis ofthese contaminants. However, it can be expected that indeveloping countries, sources will be similar to those in

developed countries although their relative importancemay be different (UNEP, 1996). A European emissioninventory is given for the by-product POPs in Table 2.This is adapted from an inventory for eight heavymetals and 15 POPs for 38 European countries carriedout by Berdowski et al. (1997). The base year is 1990and the results are a mixture of official data (submittedby 14 countries) and default emission estimates. Inaddition to their emission sources as by-products, thetable also includes sources resulting from their inten-tional use (i.e. in paints, wood preservation, othersolvent use and road paving with asphalt for PAH andin electrical equipment for PCBs). Emission estimatesare required to serve as reference points in protocols (toassess the effectiveness over time of any actions taken)and as inputs for modelling purposes. However, itshould be borne in mind that, for the European POPemission estimates given in Table 2, the uncertaintyfactors are estimated at between 2 and 5 for PAHs/PCBs and 5–20 for PCDD/Fs (Berdowski et al., 1997).

2.4.1. Dioxins and furansPolychlorinated dibenzo-para-dioxins (PCDDs) and

polychlorinated dibenzofurans (PCDFs), also referredto as dioxins and furans, respectively, are two chemi-cally similar groups of chlorinated aromatic com-pounds. The PCDD group comprises 75 and thePCDFs 135 possible congeners, 17 of which are re-ported to have potential health effects. PCDDs andPCDFs are not commercially produced but are formedunintentionally as by-products of various industrial and


combustion processes. Nearly all processes involvingthe combustion of chlorine-containing compounds re-sult in the formation of PCDDs and PCDFs. Globally,the most important source is the incineration of munic-ipal, hospital and hazardous wastes (Seys, 1997). How-ever, in Europe fuel combustion is still an importantsource of PCDD/Fs as well as the production of steeland non-ferrous metals and vehicle operation. In 1995,the PCDD/F air emissions in Europe (17 countries) areestimated to have been 6500 g I-TEQ (InternationalToxic Equivalents) per year (Quaß and Fermann,1997). This led to an average estimated deposition rateof about 5 pg I-TEQ m2/day which compares well withthe lower limit of deposition rates for rural areas inGermany.

In the human population food, particularly that ofanimal origin, is a major source of PCDDs and PCDFs(Ritter et al., 1995). PCBs and HCB are also generatedas unwanted by-products and wastes by means similarto those that generate PCDDs and PCDFs and the1990 European emissions of these are included in Table2. The major source of concern regarding PCBs is, formost countries, expected to involve leakages fromtransformers and capacitors, and PCB contaminatedwaste (UNEP, 1996) rather than as a by-product ofcombustion and industrial processes, and this isreflected in Table 2.

Improved incineration technology can reduce by-product emissions considerably by for example, follow-ing the ‘3 T’ rule (Seys, 1997):� temperature above 850°C;� time (residence) typically 2 s;� turbulence optimized by furnace geometry and sec-

ondary air.While improved incineration technology can drasti-

cally reduce combustion by-products in industrialisedcountries, rapid global industrialization, combined withpopulation growth in developing countries, has con-tributed to significant increases in the amounts of POPsgenerated from incomplete combustion, in particular,the practice of open-burning of municipal and domesticwastes common in many countries (IFCS, 1996a).Without adequately controlled combustion, the increasein quantities of chlorine-containing wastes which areincinerated could lead to further increases in the gener-ation of PCDD/Fs, PCBs and HCB.

2.4.2. Polycyclic aromatic hydrocarbonsPolycyclic aromatic hydrocarbons (PAHs) are com-

pounds containing only C and H, and are the productof any combustion process involving material contain-ing C and H (e.g. coal, oil, petrol, wood). Emissionsfrom anthropogenic activities predominate (see Table2), but some PAHs in the environment arises fromnatural combustion such as forest fires and volcanoes(Baek et al., 1991; Howsham and Jones, 1998). There

are also some minor biogenic sources (from plants,algae/phytoplankton and microorganisms) (UN-ECE,1994) and small amounts are formed by diagenesis (i.e.by slow transformation of organic materials in lakesediments). PAHs occur naturally in crude oil and forma significant component of petroleum products such assome paints, creosote (used in wood preservation) andasphalt (used for road paving) and in this sense couldbe regarded as an intentionally used class of substancesas well as being unintentionally produced by-products.PAHs are relatively reactive in the atmosphere but theycan persist long enough to become transported overlong distances to remote areas. For example, transportof PAHs to the Arctic via atmospheric suspended par-ticulate matter, clearly from coal or oil-burning sources,has been reported by Daisey et al. (1981). VariousPAHs have also been measured in remote marine atmo-spheres over the tropical north Pacific (Gagosian et al.,1981) and the Mediterranean (Sicre et al., 1987) confir-ming their long-range transport potential.

2.5. Parameters determining emission and transfer tothe atmosphere

Emissions to the atmosphere depend on a POP’sintrinsic properties, the ambient environmental condi-tions and the types of process by which it is produced,formed, applied or disposed of. Emissions can bothtake place in gaseous form and in particle-bound form,the latter being especially important during combus-tion. The intrinsic properties of a POP that largelydetermine its primary emissions in gaseous form to theatmosphere and its volatilisation from water, soil andvegetation are its vapour pressure, Henry’s-law con-stant (i.e. air–water partition coefficient) and octanol–air partition coefficient. Vapour pressure is a measureof the POP’s tendency to migrate in gaseous form fromits pure liquid or solid state but does not take intoaccount its tendency to dissolve in water or absorb toother surfaces. Henry’s law constant, being the vapourpressure divided by water solubility, does express thePOP’s tendency to migrate into air from solution inwater but it does not quantify sorption to organiccarbon or lipids. The octanol–air partition coefficient,(being the octanol–water partition coefficient dividedby Henry’s law constant) expresses the POP’s tendencyto move from air to organic media such as soils andvegetation. Experiments in which emissions of soil-ap-plied pesticides to the atmosphere were measuredshowed good correlations between the octanol–air par-tition coefficient and the percentage emission (Finizio etal., 1997a). As the values of all three parameters dependon temperature, it is important to consider this whenthese parameters are used for screening or for riskassessment purposes.


3. Transport and deposition processes

3.1. E6idence for long-range transport

Many POPs possess a combination of volatility andresistance to photolytic, biological and chemical degra-dation such that they are capable of long-range atmo-spheric transport. POPs have been measured on everycontinent and even in remote regions such as the openoceans (Hargrave et al., 1988; Iwata et al., 1993),deserts, the Arctic (Barrie et al., 1992) and the Antarc-tic (Tanabe et al., 1982; Bidleman et al., 1993) where nosignificant local sources exist. PCBs have also beenreported in human populations throughout the world(Mes, 1987). The ubiquitous distribution of these chem-icals has been known for some time. The presence ofDDTs and PCBs in the Antarctic was first reported inthe 1960s and 70s (George and Frear, 1966; Sladen etal., 1966; Risebrough and Carmignani, 1972) and con-tamination of the Arctic by several organochlorines wasevident by the 1970s (Bowes and Jonkel, 1975; Clausenand Berg, 1975; Goldberg, 1975; Barrie, 1986).

More recently it has been realized that the concentra-tions found in Polar regions are surprisingly high giventhe remoteness from sources (Wania and Mackay,1993). In the Arctic, unexpectedly high levels of certainorganochlorines were observed in seawater (Hargraveet al., 1988), precipitation (Gregor and Gummer, 1989),plankton (Bidleman et al., 1989) and wild animals(Muir et al., 1988; Norstrom et al., 1988). Weber andGoerke (1996) found concentrations of HCB in Antarc-tic fish to be as high as in North Sea fish. Levels ofPOPs found in Arctic marine ecosystems (Barrie et al.,1992) are such that there are concerns that biota,especially top predators and aboriginal peoples utilizingmarine mammals and fish in their traditional diet (Kin-loch et al., 1992), may be adversely affected by chronicexposure to these pollutants (Dewailly et al., 1989).

3.2. The global distillation/fractionation effect

One explanation for the tendency for POPs to mi-grate and deposit in Polar regions, is the ‘global distilla-tion/fractionation effect’ expounded by Wania andMackay (1993). It was suggested as long ago as 1974(Rappe, 1974) that ‘the rule of the cold wall’ mightexplain the high concentrations of some organochlori-nes in circumpolar wildlife and that pesticides used inwarm climates ‘become evaporated and transported tocool areas where they will be condensed’. The term‘global distillation’ was first coined by Goldberg (1975)in relation to the atmospheric transfer of DDT fromcontinents to oceans. This phenomenon has also beenreferred to as the ‘cold condensation’ (Ottar, 1981), the‘cold finger’ (Weschler, 1981) or the ‘cold trap’ (Rahnand Heidam, 1981) effect. More recently, Wania and

Mackay (1993) hypothesized that ‘the physico–chemi-cal properties of these chemicals, and certain factorscharacterizing cold environments, contribute to thelong-term spatial distribution patterns of organochlo-rine chemicals as much as, or even more than, therelative location of emission areas and transport path-ways. In particular the volatility of a compound andthe ambient temperature level strongly influence theobserved distribution pattern.’ In other words, theirmoderate volatility means that POPs tend to volatilizefrom tropical and temperate regions of the globe, andcondense, and then tend to remain, in colder regions.Wania and Mackay (1996) also suggest that POPsmigrate to higher latitudes in a series of relatively shortjumps, sometimes termed the ‘grasshopper effect’,whereby they migrate, rest and migrate again in tunewith seasonal temperature changes at mid-latitudes(Fig. 1). Each POP may have its own distinctive envi-ronmental condensation temperature or temperaturerange. It has been suggested that the more highlyvolatile POPs will tend to remain airborne and migratefaster and further towards the polar regions than theless volatile POPs. This should lead to individual POPsseparating out in the atmosphere along a temperatureor latitudinal gradient, according to their degree ofvolatility, in a process termed ’global fractionation’.This also implies that regions of past emission cancontinue to act as contaminant sources even after re-leases have declined or ended because of slow releasefrom accumulated chemical reservoirs.

Other evidence from both modelling (Mackay andWania, 1995) and monitoring (Muir et al., 1990; Cala-mari et al., 1991; Iwata et al., 1993; Simonich andHites, 1995) supports the above hypothesis. Also, in astudy of spatial trends and historical profiles oforganochlorine pesticides in Arctic lake sediments,Muir et al. (1995) found that their results supportedtwo of the main predictions of the cold condensationhypothesis: (i) that the more volatile organochlorineswould preferentially accumulate in polar regions and(ii) that temporal trends in their deposition would differfrom trends in temperate regions. It appears that thereis a growing scientific consensus in favour of the globaldistillation/fractionation hypothesis.

Several factors are involved in POPs’ tendency tocondense, deposit and accumulate in cooler ecosystems.Temperature-dependant physico–chemical propertiessuch as vapour pressure and solubility in water play astrong role in determining the environmental fate of allchemicals. Because of their semi-volatility, POPs canexist in the atmosphere both in the vapour phase and,in a condensed form, adsorbed onto small aerosol-sizedparticles. Knowing the physical state of a compound iscritical to the understanding of its transport and depo-sition. Atlas (1990), Bidleman et al. (1986) and othershave extensively studied the distribution of POPs be-


Fig. 1. POP global migration processes. (Reprinted with permission from Wania and Mackay (1996). Copyright (1996) American ChemicalSociety.)

tween gas and particle phases. Finizio et al. (1997b)have shown that the octanol–air partition coefficient isan excellent descriptor of this distribution. In generalterms, cooler conditions favour greater adsorption ontoatmospheric particulate matter which is then depositedmore rapidly either by dry deposition or by precipita-tion scavenging (which is more effective for particlesthan for vapours). Cooler conditions also promote en-hanced adsorption from the vapour phase onto vegeta-tion, soil, water, snow and ice at the Earth’s surface.Subsequently, accumulation will be enhanced undercooler conditions due to reduced rates of POP degrada-tion (i.e. increased persistence) and a reduced tendencyto revolatilize. POPs encompass a wide range of vapourpressures, the more volatile compounds being trans-ported further towards higher latitudes before encoun-tering temperatures sufficiently low to cause them tobecome predominantly associated with the particlephase and so subject to more rapid deposition. Thusthe global distillation/fractionation hypothesis predictsthat these chemicals will become latitudinally (and alti-tudinally) fractionated according to their volatility.Very volatile compounds may never encounter environ-mental temperatures sufficiently low to condense appre-ciably, even if they are persistent.

3.3. Uncertainties in modelling the transport anddeposition of POPs

The multi-hop nature of the environmental transport

of POPs complicates the estimation of source-receptorlinks (Wania, 1997). Attempts to relate deposition ratesto emission rates are frustrated by the lack of emissiondata, uncertainties about partitioning between gaseousand aerosol phases and a lack of data on physico–chemical properties, especially as a function of tempera-ture (Vozhzennikov et al., 1997; Atlas, 1990). There arealso few data on the effect of different surfaces ondeposition rates (Lukoyanov et al., 1992; Horstmann etal., 1997; Vozhzennikov et al., 1997) and considerableuncertainties about reaction or degradation rates ofthese substances in the atmosphere and in the media ofdeposition, including soils and water bodies (Vozhzen-nikov et al., 1997).

Despite these difficulties there is optimism that, at aregional scale and for well characterised POPs, im-provements in mass balance models are resulting in anability to estimate their concentrations in various envi-ronmental compartments, with sufficient accuracy to beof help for some regulatory purposes. Larger-scalemodels are beginning to provide valuable insights anddata which can guide and inform regulators. A reviewof these models showed that accuracy’s within factorsof 3–5 could be achieved for relatively well character-ized compounds (Cowan et al., 1995). A recent mod-elling exercise for lindane in Europe achieved a factorof 3 accuracy (Van Jaarsveld et al., 1997). Similarefforts on the series of chlorobenzenes in Canada havebeen successful in reconciling emission rates with re-


ported concentrations (MacLeod, M, and D. Mackay,An assessment of the environmental fate and exposureof benzene and the chlorobenzenes in Canada, (submit-ted for publication)). At a global scale, there has beensuccess in compiling a model of the global distributionof alpha HCH for emissions from 1947 to the presentand obtaining concentration data in the atmosphereand oceans which are consistent with monitoring data(Wania, F., D. Mackay, Y.-F. Li, T.F. Bidleman, andA. Strand, Global chemical fate of a-hexachlorocyclo-hexane. 1. Modification and evaluation of a globaldistribution model (submitted for publication); Wania,F. and D. Mackay, Global chemical fate of a-hexa-chlorocyclohexane. 2. Use of a global distributionmodel for mass balancing, source apportionment, andtrend predictions (submitted for publication)). There isa clear need for continuing efforts to obtain monitoringdata, physico–chemical properties and emissions formore POPs, and use them to further develop and testmodels at local, regional, national, continental andglobal scales.

3.4. Criteria for gauging long-range transport potential

The relevant parameters are:� Volatility—measured by vapour pressure;� Persistence in air or water—measured by half-lives;� Documented evidence (i.e. monitoring data) that the

substance undergoes long-range transport and isfound in remote regions at relevant concentrations.

3.4.1. VolatilityIn the CLRTAP POPs protocol a vapour pressure

criterion of B1000 Pa is suggested. The rationale forthis is partly that it excludes volatile chemicals, such asCFCs, which are the subject of other protocols. It alsoreflects the view that volatile substances are unlikely topartition into terrestrial or aquatic systems. This crite-rion excludes relatively few chemicals and probably, itcould be eliminated with few adverse consequences. Itcan be argued that, for a substance to be transported inthe atmosphere, it should have a vapour pressure ex-ceeding some minimum value. However, it would bedifficult to specify such a criterion because of thepossibility of the transport of involatile compounds inan aerosol phase and the complication of selectionbetween solid and liquid vapour pressures. It is alsodifficult to measure very low vapour pressures. In thefuture, it may be preferable to specify minimum andmaximum limits to the octanol–air partition coefficient.This coefficient can be calculated as the ratio of oc-tanol–water and air–water partition coefficient and isbelieved to give an excellent characterization of parti-tioning from air to aerosols, soils and vegetation. It isstill a fairly recently developed concept and few mea-surements are available, but is could prove valuable in

future assessments of the relative potential for atmo-spheric transport.

3.4.2. Persistence in air or waterA chemical must have a sufficiently long atmospheric

residence time in order to be subject to long-rangeatmospheric transport. The two major processes con-trolling atmospheric residence time are degradation rateand rates of deposition/recycling. Important parametersdetermining the degradation rate of POPs in the atmo-sphere are temperature and the reaction rate constantfor reaction with the OH-radical. The overall degrada-tion rate is generally estimated by a half-life value. Anatmospheric half-life greater than 2 days is considerednecessary for long-range atmospheric transport in theCLRTAP POPs protocol.

The partitioning of a POP between the gas andparticle phases affects its residence time with highervolatility tending to confer longer residence times. Thevapour pressure of the POP and the ambient air tem-perature are therefore important parameters determin-ing its atmospheric residence time. Deposition will takeplace either as wet or dry deposition. Wet depositionincludes both POPs dissolved in the precipitation andPOPs bound to particles. Henry’s law constant (H) isan important parameter for the partition of POP be-tween gas and water phases. The gas/surface (water,soil, vegetation) exchange process has also been shownto be significant for POPs, important parameters beingvapour pressure, H, and the octanol–water and oc-tanol–air partition coefficients.

Under the proposed UNEP global POPs initiative,long-range transboundary transport by other means,such as in ocean currents and rivers, will also have tobe considered. However, it should be noted that aparticular half-life in water appropriate for assessingoceanic transport potential would be inadequate formore rapid, riverine transport.

4. Bioaccumulation

4.1. Definitions

Bioconcentration and the bioconcentration factor(BCF) refer to the uptake of a chemical by an aquaticorganism from water. Bioaccumulation refers to uptakefrom both water and food and the bioaccumulationfactor (BAF) is the ratio of organism and water concen-trations. Biomagnification factors are usually the ratioof the concentration in the predator to that in the prey,possibly on a lipid-adjusted basis. Not all of the chemi-cal present in the water is necessarily available becauseof sorption to suspended and dissolved matter, usuallyorganic in nature. This phenomenon is referred to asbioavailability. The most frequently employed descrip-


Table 3Bioaccumulation and biomagnification of organochlorines in Arctic marine food chains (after Muir et al., 1992 using data from a variety ofpublished sources

Fish to seal Fish to beluga Fish to bird Seal to bearWater to zoo-plankton Water to fish

—4.6 —Toxaphene 0.386.7×104 2.6×106

0.7 9.3HCH 9.3×103 3.0×104 1.5 1.46.60.339.4Chlordane 7.32.6×106 5.3×107

8.5 0.3DDT ]4.3×105 ]1.4×107 20.2 9.85.8 7.4PCB* 1.9×106 4.8×107 8.8 8.0

3.0 7.5HCB 1.2×105 9.6×106 15.60.2

a PCBs as Aroclor equivalents.Organism to organism values are on a lipid /lipid basis).

tor of bioconcentration and bioaccumulation for non-polar organic compounds is the octanol–water parti-tion coefficient, which quantifies hydrophobicity as adeterminant of partitioning from water to lipid phases.

4.2. Mechanisms of bioaccumulation

Because of their low solubility in water and theirresistance to chemical and metabolic degradation, mostPOPs are eliminated from organisms very slowly (UN-ECE, 1994). As a consequence of this persistence, POPscan accumulate to relatively high levels in biota even atlow environmental exposures. Phytoplankton can sorbdissolved DDT, HCB and PCB directly from the waterand it has been suggested that DDT may be accumu-lated directly through gill membranes in Arctic char(Swain et al., 1992). However, in the aquatic environ-ment the primary route of initial entry into the foodchain is through active uptake of POP-contaminatedparticulate matter by filter-feeders and plankton(Thomann et al., 1992) and at higher trophic levelsdietary uptake is more important than direct absorp-tion (UN-ECE, 1994). Deposition and degradation pro-cesses and the bioavailability of POPs in terrestrialecosystems are less well understood. However, it isknown that POPs can be absorbed through plant sur-faces; pine needles have been found to be useful indica-tors of atmospheric contamination by organochlorines(Eriksson et al., 1989). For HCB, HCH and toxaphene,the indications are that the air/plant/animal contami-nant pathway is the major route taken by these com-pounds into the arctic terrestrial food chain (Thomas etal., 1992).

The resistance of POPs to chemical and metabolicdegradation means that they can become more concen-trated the further they move up through food chains(i.e. they are subject to biomagnification). Biomagnifi-cation can lead to concentrations in top predatorsmany orders of magnitude higher than in the environ-ment. This is especially true of aquatic food chainswhich tend to be more complex and longer than terres-

trial food chains. Table 3 shows that, for Arctic marinefood chains, bioaccumulation factors (BAFs) from wa-ter to higher trophic level predators (seal, beluga,seabirds and bears) can be in the order of 107 fortoxaphene and HCB and as high as 109 for PCBs.

4.3. Predicting the bioaccumulation potential of a POP

The level of bioaccumulation in a target speciesdepends on the food chain structure of the particularecosystem in which it lives. For example, Baleen whaleswhich skim the water for plankton are less contami-nated than predatory species living in the same region(Ballschmitter, 1996). Even different populations of thesame species, living in areas of equivalent ambientcontamination, can contain very different levels ofPOPs simply because of different feeding patterns. Forexample, it was discovered that in one Canadian walruspopulation, where levels of contamination with PCB,DDT, toxaphene and chlordane were particularly high,the animals were feeding on seals as well as molluscs(Segstro et al., 1993). They were, therefore, feeding at ahigher trophic level than other less contaminated walruspopulations which largely subsisted on molluscs.

Biomagnification in a particular ecosystem can bestudied using stable isotope analysis with the ratio of15N abundance relative to 14N in an organism increas-ing on average 3–5 parts per thousand (‰) comparedwith its source of dietary nitrogen (Peterson and Fry,1987) due to the preferential excretion of the lighterisotope during metabolic processes (Gaebler et al.,1966). Using this technique, Kidd et al. (1995) showedthat biomagnification through an unusually long foodchain in Lake Laberge, Canada was the reason thatlake trout and burbot contained amounts of toxapheneand other lipophilic contaminants several times greaterthan levels found in the same species inhabiting someother subarctic Canadian lakes. The bioaccumulationpotential of a substance depends on the differentmetabolic characteristics of different species. This canresult in large differences in biotic persistence for aparticular POP. For example, a particular dioxin can


have a half-life of a few weeks in rats whilst in humansit is 7–9 years (Ballschmitter, 1996). Also, differences inthe ability of a single animal species to metabolisedifferent POPs, or even the different congeners of thesame group of POPs, often hinder attempts to predictbiomagnification (Muir et al., 1988). For risk assess-ment purposes it can be concluded that it is difficult topredict the actual level of exposure of a receptor organ-ism, for a given concentration of POP in its ambientenvironment, without having detailed knowledge of thespecies’ position in the ecosystem food web and of thePOP’s biotic persistence throughout the relevant foodchain. Despite these difficulties there have been suc-cesses in applying bioaccumulation models to singleorganisms and even food chains when there is adequateunderstanding of the properties of the POP, itsbioavailability and the characteristics of the organismsand their diets.

4.4. Bioaccumulation criteria

The primary criterion for bioaccumulation for allchemicals is the bioaccumulation factor (BAF) or infish, the bioconcentration factor (BCF). A secondarycriterion for non-polar, hydrophobic organic chemicalsis the octanol–water partition coefficient, if the molecu-lar weight of the substance is less than some criticalvalue, variously estimated to lie between 600 and 1200,and if the substance is not metabolized.

5. Persistence

The simplest approach is to set criteria for half-livesin the four primary media of air, water, soil andsediment. For example, the following criteria have beensuggested for the CLRTAP POPs protocol:

Half-life in air \2 days;Half-life in water \2 months;Half-life in soil or sediments \6 months.Although these criteria appear simple and straight-

forward there are likely to be severe difficulties whenapplying them:� Unlike a radionuclide, the half-life of a POP depends

on its environment, e.g. the temperature, redoxstatus, incidence of sunlight, the presence of reactivechemical species and the nature of microbial commu-nities. A specific POP thus experiences a spectrum ofhalf lives with variations in time and space. Applyinga sharp cut off or a ‘bright line’ is thus likely to bedifficult and controversial.

� Although there are established test procedures forbiodegradability and for assessing pesticide fate insoils, no standard test protocols exist for measuringhalf-lives, in the laboratory or the environment, inwhich all relevant processes in these media are in-

cluded. There are also difficulties in translating labo-ratory results to the environment.

� The overall persistence of a chemical in the environ-ment depends on how it is emitted to the environ-ment (i.e. to air, water, or soil) and on how itsubsequently migrates between media. The implica-tion is that a substance may be quite short-lived ifdischarged to air, but long lived if it is discharged towater. Furthermore, a long half-life in a mediummay be relatively inconsequential if the substance isnot emitted to that medium or is unlikely to transferto it. For example, an accurate half-life for reactionin air may not be needed for a relatively involatilechemical which is unlikely to evaporate into theatmosphere.

� It should be appreciated that if a criterion of say 6months is used as a half-life, it will be necessary toobserve the chemical concentration decrease for atleast two half lives or 1 year. Under environmentalconditions, it is inevitable that conditions such astemperature will change substantially during thatperiod and it will be difficult to obtain reproducibleresults.

� In determining the half-life for a medium, all reac-tion mechanisms, including biodegradation, hydroly-sis, photolysis etc., should be considered but notlosses by transport to other media (e.g. by sedimen-tation or evaporation). If significant migration be-tween media occurs, the residence time in eachmedium should also be considered in order to esti-mate the overall persistence in the environment.Despite these difficulties, there remains a need to

characterise persistence in order that the highest prior-ity can be applied to the most persistent substances.The general magnitudes of the suggested half-lives arejudged correct, at least for initial screening purposes.More effort is needed to develop experimental proto-cols and predictive methods, and to gather data onobserved environmental degradation rates. Where datais not available, expert judgement should be used. If theweight-of-evidence is not convincing then a substanceshould be initially assumed to meet the criteria basedon the precautionary principle.

6. Metabolism and mechanisms of action in biota

The concept of biotic persistence implies that com-pounds undergoing biomagnification are unlikely to bebiotransformed at a high rate in living cells. It might beexpected that POPs would not tend to exert toxicityfollowing metabolic transformation and indeed, hightoxicity is generally associated with metabolically stablecompounds such as PCDD/Fs and co-planar PCBs.Research during recent years has demonstrated a num-ber of important exceptions to this general rule with


methylsulphonyl-containing POP metabolites (e.g. ofPCBs and DDT) having being found in mothers milk,as well as in the tissues of marine and terrestrial mam-mals and birds. Recently, a large number of hydroxy-lated metabolites derived from chlorinated andbrominated aromatic compounds have been demon-strated in human blood plasma as well as in the bloodof wild mammals (seals), birds (albatrosses) and fish(salmon). These metabolites exhibit a number of spe-cific effects that add to the profile of activities of theirparent compounds (Brouwer et al., 1998). In this sec-tion, the receptor-based mechanisms of action of somePOPs, and the formation and effects of their metabo-lites, are discussed.

The rate of formation of metabolites is highly depen-dent on the degree and position of chlorination. Ingeneral, metabolite formation decreases with increasingchlorination, e.g. the half-life of 3,3%,4,4%-tetrachloro-biphenyl (3,3%,4,4%-TeCB) is 2 days in rats, while2,2%,4,4%,5,5%-hexachlorobiphenyl shows a lifetime elimi-nation of only 20% of the dose given in the samespecies. In addition, the position of the chlorine-sub-stituents is an important factor. Hydroxylation tends totake place on meta and para-positions of aromaticrings. Therefore, POP congeners with chlorine substitu-tion on meta and para positions show much lower ratesof metabolism, e.g. the stable congeners of PCDD andPCDFs are all 2,3,7,8-substituted congeners. Moreover,the rate of metabolism is highly species dependent. Ingeneral, lower organisms like crustacea and fish have alower metabolizing capacity than mammals. These spe-cies and congener-dependent differences in rates ofmetabolism will result in considerable differences in thecongener patterns of parent compounds and the pres-ence and amount of metabolites formed in differentspecies. This may have important implications for thetoxic potency as well as the profile of effects of POPs.In fact, there is a wide species difference in sensitivityand toxic profile for PCBs, PCDDs and PCDFs.

One important molecular mechanism of action ofPCDDs, PCDFs and PCBs appears to be receptor-based, involving the so-called Aryl hydrocarbon recep-tor (AhR). When entering cells, compounds like dioxinsbind with high affinity to the cytosolic AhR protein,which then undergoes a process of activation andmoves to the nucleus where the liganded-AhR is boundto specific elements (dioxin response elements (DRE))on the DNA. This results in an increased transcriptionof the genes that possess a DRE-element in their up-stream control units for expression, such as theCYP1A1/2 genes. Consequently, there is a highly in-creased de novo synthesis of CYP1A1/2 enzymes andother proteins that may be involved in other aspects ofthe toxic profile. Since AhR-binding affinity, enzymeinduction and toxic potency correlate well over a widerange of different congeners it is nowadays widely

accepted that the AhR plays a major role in the onsetof the toxic profile of many POPs. In fact, CYP1A1/2induction is widely used nowadays as a biomarker forexposure to dioxin-like POPs. Furthermore, the AhR-theory is used as the basis for the toxic equivalencyfactor (TEF) concept (Safe, 1994) which uses the rela-tive potencies of individual congeners multiplied bytheir concentration, summed up over all congeners in amixture to give rise to a sum-toxic potency, the toxicequivalent (TEQ).

6.1. Effects of hydroxylated POP metabolites

Hydroxylated metabolites of PCBs, PCDDs andPCDFs, as well as of chlorinated benzenes have beenfound in blood plasma and in several tissues (liver,brain) in experimental animals, as well as in plasmafrom free-ranging marine mammals and in human sub-jects (Bergman et al., 1994). In laboratory animals, ithas been observed that several of the hydroxy-metabo-lites of PCBs have a high tendency to cross the placen-tal barrier, so entering the foetal compartment, and tocross the blood-brain barrier. (Darnerud et al., 1996;Morse et al., 1996; Brouwer et al., 1998).

Hydroxy–PCBs exhibit their own specific profile ofeffects, distinct from the effects caused by the parentPCBs. These effects include, direct interference in en-docrine systems, including the thyroid system, the estro-gen system as well as the retinoid system (Brouwer andVan den Berg, 1986; Jansen et al., 1993; Brouwer et al.,1998). For example, hydroxy–PCBs resemble the struc-ture of the thyroid hormone, thyroxine (T4), resultingin competitive displacement of the natural thyroid hor-mone, T4 from its major binding proteins, such asplasma transthyretin, the T4-converting enzymes, 5%-deiodinase, and T4-sulfotransferases. Some hydroxy–PCBs show a weak estrogenic potential acting via theestrogen receptor; other hydroxy–PCBs however, werefound to be mainly anti-estrogenic. Hydroxy–PCBs arepotent uncouplers of mitochondrial oxidative phospho-rylation. Furthermore, they inhibit intercellular com-munication which may result in uncontrolled cellulardifferentiation and proliferation, tumour promotionand alterations in developmental processes.

6.2. Formation and effects of methylsulphonemetabolites

The methylsulphones are formed in the complexpathway involving entero-hepatic circulation and intes-tinal microbial metabolism of glutathione (GSH) conju-gated xenobiotics. A number of methylsulphonesderived from PCB, DDT and chlorinated benzenes areselectively accumulated and retained in specific celltypes and tissues in mammals and birds (Brandt andBergman, 1987). A number of PCB-derived methylsul-


phones become selectively bound in high concentrationsin non-ciliated bronchiolat (Clara) cells following asso-ciation to a uteroglobin-like PCB-binding secretoryprotein residing in these cells (Brandt et al., 1985; Lundet al., 1985). The methylsulphone-protein complex sub-sequently becomes secreted into the airways. A similarPCB-binding protein present in the human airways maybe of importance for the development of chronic res-piratory distress observed in victims of the Yusho PCBintoxication incident in Japan in 1968.

Methylsulphonyl–DDE is a persistent DDT metabo-lite present in human milk and the blubber of Balticgrey seal (Jensen and Jansson, 1976). In mice, thiscompound is covalently bound following a cytochromeP450-catalysed activation in the adrenal zona fascicu-lata, i.e. the site of glucocorticosteroid hormonal syn-thesis in mammals. The activating P450 form (P45011B1)controls the final step in the glucocorticosteroid synthe-sis and is specifically expressed in the zona fasciculatacells. Consequently, methylsulphonyl–DDE is a highlypotent toxicant that induces degeneration and/or necro-sis of the zona fasciculata cells following a single doseas low as 3 mg/kg body weight. Decreased plasmacorticosterone levels are observed both in adult animalsand in sucking pups following dosing of the nursingdam (Lund et al., 1988; Brandt et al., 1992). Thesefindings are of particular interest because adrenocorti-cal hyperplasia was a cardinal finding in Baltic sealsexposed to high body-burdens of POPs (Jonsson et al.,1991, 1992).

There seem to be strict structural requirements forthe toxicity of methylsulphonyl metabolites. While onemethylsulphonyl–dichlorobenzene isomer induced irre-versible toxicity following a single dose, another isomercompletely lacked toxicity despite displaying a corre-sponding level of irreversible binding in the olfactorymucosa.

7. Effects

7.1. Introduction

The acute effects of intoxication on animals (fromlaboratory studies, and accidental spills) and on hu-mans (from accidental consumption of contaminatedfood, e.g. the HCB poisoning in Turkey, or accidentalrelease of dioxin as happened at Seveso, Italy in 1976)have been fairly well documented for certain POPs. Thelong-term effects of high-level accidental releases arestill not very well known. Recently, Mocarelli et al.(1996) described a female-skewed sex ratio in birthsabout eight years after the Seveso accident which canbe directly related to dioxin exposure. Acute and

chronic occupational exposure to POPs is of concern insome developing countries where they continue to beused in tropical agriculture.

The often complex and subtle effects of chronic,low-level environmental exposure to POPs are less wellunderstood. In the environment, the universal exposureof organisms to low levels of a wide range of chemicalcontaminants (and possibly, to stresses of a non-chemi-cal nature) makes it extremely difficult to ascribe anobserved effect to any particular one of them. There isalso the possibility that, in the environment, toxic sub-stances in combination may act additively, antagonisti-cally or synergistically. There are experimental animaldata indicating the existence of interactive effects forcomplex mixtures of POPs. For example, the PCBmixture Aroclor 1254 antagonises the toxicity of 2,3,7,8TCDD with respect to hepato, immuno and reprotoxic-ity (Davis and Safe, 1989). On the other hand, VanBirgelen et al. (1996) found that, while PCB-153 alonedid not result in porphyrin accumulation in rats, co-ad-ministration of PCB-153 with dioxin revealed a strongsynergistic effect. In addition, when comparing theSeveso (2,3,7,8 TCDD) and Yusho/Yucheng (PCBs+dibenzofurans) accidents, a much wider range of effectsis observed in the PCB poisoning than in the TCDDintoxication. This suggests that interactive effects ofPCBs and PCDD/Fs can also occur in humans.

Attempts to link cause with effect are further compli-cated because the effects of POP exposure vary consid-erably depending on species, age, gender and the level,duration and timing (relative to the organism’s life-cy-cle) of exposure (Swain et al., 1992) and because theremay be a considerable delay (up to two or three gener-ations) between POPs exposure and the onset of effects(Murrey et al., 1979). A POP may also degrade to forma more (or less) toxic product requiring knowledge ofthe various possible transformations that could occur.

There is, however, experimental evidence for somePOPs that cumulative low level exposures in animalsmay be associated with chronic sub-lethal effects suchas immunotoxicity, dermal effects, impairment of re-productive performance and carcinogenesis (Ritter etal., 1995). Brouwer et al. (1995) have described the widerange of neurobehavioural, reproductive and endocrinealterations observed in experimental animals followingin utero and lactational exposure to PCBs and PCDD/Fs. There are also subtle changes observable in neu-rodevelopmental and thyroid hormone parameters inhuman infants at background human body burdens.Consequently, in assessing the effects linked to thepresence of POPs in the environment, the full range oftoxic endpoints should be considered including themore subtle and complex chronic endpoints describedbelow.


7.2. Chronic effects on experimental animals andwildlife

7.2.1. Immune system effectsImmunosuppression may be one of the most sensitive

and relevant environmental threats posed by POPs(Han and Stone, 1997). In numerous experiments withanimals, it has been shown that various POPs cansuppress immune system function. Subchronic levels ofDDT impaired the immune response of mice (Banerjeeet al., 1986; Banerjee, 1987), aldrin and dieldrin reducedmouse resistance to viral infection, chlordane and hep-tachlor were found to affect animals’ developing im-mune systems and lindane affected macrophage activityin vitro as well as reducing experimental animals’ resis-tance to infection (Barnett and Rodgers, 1994). Labora-tory studies have also shown that immune systemfunction can be impaired by PCDDs (Holsapple et al.,1991), PCBs (Tryphonas et al., 1991) and HCB (Barnettet al., 1987).

Wildlife studies also reveal evidence for the im-munotoxic effects of POPs. Herring gull and Caspiantern chicks in the Great Lakes region were found tohave smaller thymuses and less T-cell activity thegreater their exposure to certain organochlorines (Gras-man, 1995). Autopsies on dead beluga whales from theSt. Lawrence Seaway, Canada (Martineau et al., 1987,1988) found high tissue concentrations of organochlo-rine pesticides and PCBs and severe bacterial infectionssuggesting immunosuppression. De Guise et al. (1994,1995) also found more frequent and severe tumours inthe St Lawrence beluga whales compared with whalepopulations elsewhere and concluded that exposure tocarcinogenic compounds and decreased resistance tothe development of tumours could both have con-tributed to this. Researchers who assayed blood takenfrom live bottlenose dolphins off the Florida coastfound that decreased T-cell lymphocyte proliferativeresponse was highly correlated with increased levels ofbioaccumulative organochlorines, the dolphins also ex-hibiting infections suggestive of immune dysfunction(Lahvis et al., 1993). In the early 1990s, a massivedie-off of striped dolphins in the Mediterranean Seawas attributed to infection by a morbillivirus similar tothat causing distemper in carnivores (Aguilar andRaga, 1993). Diseased dolphins were found to carryconcentrations of several organochlorines (in particularPCBs) between two and three times the levels com-monly found in the healthy population and it has beensuggested that the immunodepressive capacity of thePCBs may have been a contributory factor (Borrell andAguilar, 1991). The morbillivirus was also reported tobe responsible for the unusual die-offs of NorthernEuropean harbour seals in the Baltic and the North Seain 1987 and 1988 (Kennedy et al., 1988; Osterhaus andVedder, 1988; Dietz, et al., 1989) and of Baikal seals in

1988 (Grachev et al., 1989). Brouwer et al. (1989) foundthat consumption of PCB-contaminated fish from theDutch Waddensee led to vitamin A and thyroid hor-mone deficiency in the common seal, both of whichcould result in increased susceptibility to microbialinfections. Further evidence that certain organochlori-nes may have compromised the seals’ immune systemsand so contributed to the die-offs came from a prospec-tive immunologic study on harbour seals (De Swart etal., 1993, 1994, 1995; Ross et al., 1995). Seal pups werefed uncontaminated fish for a 1 year acclimatizationperiod and then split into two groups. One group wasfed with herring caught in the polluted Baltic Sea whilstthe other group was fed herring from the relativelyunpolluted waters around Iceland. The dietary intakeof organochlorine pesticides and PCBs of seals fedBaltic fish was several-fold higher than that of thecontrols and led to a 10-fold higher concentration intheir blubber. The group of seals eating Baltic herringshowed significantly reduced natural killer cell activityand reduced T-cell proliferative response, antibody re-sponse and neutrophil levels. This group ended up withimmune responses estimated to be three times weakerthan those of the control group and increases in oppor-tunistic infections indicative of immunosuppression.This was the first demonstration of immunosuppressionin mammals as a result of exposure to contaminants atambient levels found in the environment (Osterhaus etal., 1995).

7.2.2. Effects on the endocrine systemThe potential effects of so-called endocrine disrupt-

ing compounds (EDCs) on the endocrine system haverecently gained much attention. It has been suggestedthat many different adverse health endpoints in wildlife,animals and in humans are or may be, associated withexposure to a variety of EDCs, both anthropogenic andnatural in origin. Several agencies have proposed defin-itions to describe what is actually meant by an en-docrine disrupting chemical. The definition proposed bya workshop held at Weybridge, UK in 1996 (EuropeanCommission, 1996) was as follows: ‘An endocrine dis-rupter is defined as an exogenous substance that causesadverse health effects in an intact organism, or itsprogeny, consequent to changes in endocrine function’.

Several substances classified as POPs have been indi-cated as being potential endocrine disrupters (Toppariet al., 1996). The evidence is partially based on in vitrostudies, investigating the potential of a chemical to bindwith for example, the estrogen receptor and to activatethe receptor pathway in terms of altered gene expres-sion. For example, several hydroxy–PCBs were testedin in vitro assays for comparative binding to the estro-gen receptor. The hydroxy–PCBs all showed someestrogenic receptor binding potency, the most potentone, 4-OH-2%,4%,5%-triCB was about 80-times less potent


than the synthetic estrogen, estradiol (Golden et al.,1998). In functional in vitro assays, like proliferation ofestrogen-sensitive MCF-7 cells, most of the hydroxy–PCBs tested were found to be anti-estrogenic (Moore,et al., 1997). For DDT, both o,p %-DDT, o,p %-DDD ando,p %-DDE were very weakly estrogenic in in vitro hu-man estrogen receptor assays.

There are also experimental animal data on hor-monal changes and alterations in physiological parame-ters. For example, the PCB mixtures Aroclor1221,1232, 1242 and 1248 exhibited some degree ofestrogenic activity in an in vivo uterotrophic assay,roughly six orders of magnitude lower than estradiol.There were also some changes in the LH (LuteinisingHormone) and FSH (Follicle Stimulating Hormone)secretion by the pituitary. For hydroxy–PCBs, a corre-lation was found between in vitro estrogen receptorbinding and in vivo uterotrophic activity. Also, o,p %-DDT and o,p %-DDE were found to be estrogenic in invivo uterotrophic assays; about 2500 and 40000 timesless potent than estradiol. For hydroxy–PCBs as wellas parent PCBs it is well known that thyroid hormonelevels are affected in experimental animals (Brouwer etal., 1998). The interaction of the PCBs takes place atmultiple levels of the thyroid hormone system, includ-ing the thyroid gland, the pituitary-thyroid feedbacksystem, transport of thyroid hormones and enzymaticconversion of thyroid hormones. This also occurs infetuses born from dams exposed to PCBs. Althoughboth hormonal changes and physiological/developmen-tal effects can be induced following exposure to certainPOPs, and it is tempting to suggest that these arelinked, there is actually little or no conclusive evidencethat this link exists.

There are also a number of wildlife cases where a linkbetween a chemical exposure, its endocrine disruptingeffect and observed adverse health outcomes is sug-gested. For example, associations between elevatedresidues of PCBs and DDTs and reduced nesting suc-cess, increased chick and embryo mortality or reduc-tions in population have been reported for Europeanfish-eating birds (Colborn, 1991). Delayed or abnormalsexual differentiation and/or development have beensuggested to be associated with population declines ofalligators and the presence of organochlorine pesticidesin lakes in Central Florida (Guillette et al., 1994).Environmental exposure to DDE and PCBs has beenimplicated in various reproductive problems (e.g. lowsperm counts, high levels of sperm abnormalities andextremely high incidence of undescended testicles) inthe endangered Florida Panther population (Facemireet al., 1995). Concentrations of certain organochlorinepesticides, PCBs, and PCDD/Fs were negatively corre-lated with testis size and weight, and with the length ofthe baculum (the penile bone), in young male ottersfrom the Columbia River, USA (Henny and Grove,

1996). The recent Weybridge workshop on endocrinedisrupters concluded that ‘some cases exist within theEU area where adverse endocrine effects, or reproduc-tive toxicity, in birds and mammals coincide with highlevels of anthropogenic chemicals which have beenshown to have endocrine disrupting properties in sometest systems’ (European Commission, 1996).

7.2.3. Reproducti6e and de6elopmental effectsIn a literature review by Wren (1991) many labora-

tory experiments with mink and ferrets are describedwhich demonstrate a direct cause and effect relation-ship between PCB exposure and immune dysfunction,reproductive failure, enzyme induction, increased kitmortality, deformities, organ enlargements and adultmortality. Declines in European otter populations havealso been attributed to environmental exposure toPCBs and other organochlorines (Mason, 1989; Olssonand Sandegren, 1991). There are also many well-knownexamples of bird populations that have been affected byPOPs exposure with decreased or retarded egg produc-tion, increased embryo mortality, eggshell-thinning, em-bryonic deformities, growth retardation and reducedegg hatchability being among the effects reported (Hanand Stone, 1997). For example, Gilbertson et al. (1991)related embryo deformities and egg mortality in GreatLakes fish-eating birds to organochlorine contaminants.Fox (1992) described abnormalities in Western andHerring gulls in the Great Lakes including increasedfemale–female pairings, and altered nest defence andincubation behaviour resulting from sex ratios beingskewed towards females. More recently, pesticides andother synthetic chemicals have been linked with bizarredeformities (e.g. extra legs growing from the abdomenor neck) in frogs across Minnesota and Wisconsin andin the St. Lawrence River Valley in Quebec (Ouellet etal., 1997) although other possible causes are also beingconsidered.

Moreover, reproductive failure in the common sealpopulation inhabiting the western part of the DutchWadden Sea was attributed to PCBs after feeding ex-periments demonstrated that a diet of naturally PCB-contaminated fish from this sea had a detrimental effecton the seals’ reproduction (Reijnders, 1986). Mac andEdsall (1991) demonstrated reduced hatching successand diminished fry survival in lake trout eggs experi-mentally exposed to PCBs and DDTs and found therewas a gradual increase in hatchability and fry survivalof Lake Michigan trout since 1984 following a reduc-tion in organochlorine levels in the Great Lakes. In theBaltic in recent years, researchers have seen extremedisturbances of reproduction in several places andamong several fish species including perch, burbot, codand salmon. Although the mechanisms for these distur-bances are unproven, it is probable that one or morePOPs are involved (Alsberg et al., 1993).


A clear cause and effect relationship with POPs andreproductive failure has been established for mink.Ranch mink exhibit reproductive failure when fedGreat Lakes fish containing between 0.3 and 0.5 mg/gPCBs (Aulerich and Ringer, 1977) and tissue levels ofPCBs in mink above 50 mg/g produced reproductivefailure in captivity (Jensen et al., 1977). Lopez-Martınet al. (1994), in a study of the threatened Europeanmink in Northern Spain, found PCB body burdens tobe well over 50 mg/g (with means of 118 mg/g in muscleand 123 mg/g in liver) and suggested that these pollu-tants could be partly responsible for their decline. Thereproductive effects and abnormalities observed inwildlife are consistent with similar effects induced byvarious POPs in laboratory experiments on animals.Fry and Toone (1981) induced feminization in maleCalifornia gulls by injecting eggs with DDT in amountscomparable to those found in seabird eggs in southernCalifornia in the late 1960s. Chlordane disturbed sper-matogenesis and damaged the testes of mice (Balash etal., 1987). HCH fed to male weanling rats decreasedtestis weights at doses \50 mg/kg with testicular atro-phy resulting from a dose of 250 mg/kg (Velsen et al.,1986). Peterson et al. (1992) describe experiments inwhich in utero and lactational exposure to a dioxin(2,3,7,8-TCDD) delayed testis descent, reduced theweight of sex organs, inhibited spermatogenesis anddemasculinized male rat sexual behaviour. Gray (1992)also describes the adverse impacts on rodent neuroen-docrine and hypothalamic sex differentiation and theproduction of anatomical malformations resulting fromperinatal exposure to estrogenic pesticides, PCDDs andPCBs.

7.2.4. CarcinogenesisThere is convincing evidence from laboratory experi-

ments that several POPs can have genotoxic effects oract as tumour promoters (Perera, 1981; Nagayama etal., 1992; Wolfe and Marquardt, 1993). Stranded car-casses of St. Lawrence beluga whales were found tohave a high incidence of tumours and contained ele-vated levels of PCBs, mirex, chlordane and toxaphene(UN-ECE, 1994). Only in the case of PAHs (some ofwhich, particularly benzo(a)pyrene, are highly carcino-genic) have direct links been established between expo-sure and increased cancer rates in wildlife. For example,the incidence of dermal lesions, tumours and papillo-mas in the brown bullhead (a bottom-feeding fish fromthe Great Lakes) was significantly elevated in areaswith high PAH levels in the sediments (Government ofCanada, 1991).

7.2.5. NeurotoxicityDevelopmental toxicology is a rapidly evolving area

of environmental toxicology. Much of this progressstems from the observation that there are periods of

enhanced sensitivity to chemical toxicity during theearly life-stages of organisms, when both structuralmalformations and persistent functional changes can beinduced. With regard to neurotoxicity, a period ofextreme sensitivity seems to occur when the developingbrain enters a phase of rapid growth, i.e. the so-called‘brain growth spurt’. Depending on the species exam-ined, this period occurs either during mid-late gestation,around term or in the early post-natal period. In miceand rats, this sensitive period occurs around day 10post partum, while in guinea pigs the period occurs inmid-late gestation. In humans, the period begins duringthe third trimester of pregnancy and continues through-out the first 2 years of life.

Studying a number of POPs and non-persistent neu-rotoxicants, Eriksson et al. have shown that a singlelow-dose exposure at day 10 post partum may result inpersistent changes in spontaneous behaviour (hyperac-tivity) and learning ability at adult age (Eriksson, 1992;Eriksson and Fredriksson, 1996; Eriksson, 1997). Thesebehavioural changes were correlated with persistentchanges in the expression of cholinergic receptor sub-types. However, following dosing at day 3 or 20 postpartum, no corresponding effects were observed atcomparative doses. As mentioned above, these develop-mental changes are not restricted to POPs, but can alsobe induced by certain non-persistent organic chemicalssuch as pyrethroid, paraquat and nicotine. Interest-ingly, several persistent environmental pollutants suchas DDT (a neurotoxicant by design) and several PCBsinduce this type of neurotoxicity, while other PCBsseem to lack activity at corresponding doses. For in-stance, when PCB 52 was applied as a single dose (1.4mmol/kg; 0.4 mg/kg) on day 10 post partum, hyperac-tivity was observed at adult age. Preliminary resultsalso show that a polybrominated diphenyl ether (PBDE99) induces behavioural toxicity at equimolar doses (1.4mmol/kg; 0.8 mg/kg).

7.3. Chronic effects on humans

The weight of scientific evidence suggests that somePOPs have the potential to cause significant adverseeffects on human health both at a local level and,through long-range transport, at regional and globallevels (Ritter et al., 1995). For some POPs, especiallythe pesticides, occupational and accidental high-levelexposure is of concern for both acute and chronicexposure. This is particularly true in the developingworld where inadequate safety and hygiene practicesare the norm in applying, formulating, storing, trans-porting and manufacturing pesticides (Repetto andBaliga, 1996). Pesticide warning labels are often printedincorrectly or in the wrong language and many usersare illiterate. In tropical climates, sprayers rarely wearthe uncomfortable and costly protective clothing. For


Table 4Pesticide concentrations in adipose tissues in Greenland and other countries (after Mulvad et al., 1996)

Canada 1983 USA 1983–1984 Finland 1983Greenland (N=42) 1993

2.1 1.2PCB (mg/kg) 0.315.81670 33026304450DDT (mg/kg)

752 10 31 20HCB (mg/kg)

example, in the Northern Philippines farmers sprayingpesticides wore masks in only one quarter of sprayingoperations and never wore gloves or boots (Alba,1988). Agricultural field workers rarely observe there-entry period, the time required between spraying andother fieldwork. Weeders as well as children and otherhousehold members in or near newly sprayed fields arealso directly exposed to pesticides (Rola and Pingali,1993).

In any discussion of substitutes for POP pesticides,the greater acute toxicity of many non-POP pesticidesshould be borne in mind. For example, in the early1970s the introduction of the neurotoxic organophos-phates and carbamates for disease vector control, asalternatives to DDT, resulted in a considerable increasein acute poisoning cases, both accidental and occupa-tional (WHO, 1996).

As noted for wildlife, it is extremely difficult toestablish cause and effect relationships between humanexposure to low levels of a POP in the environment andparticular adverse health effects, not least because ofthe broad range of chemicals to which humans areexposed at any one time. The contamination of food,including breast milk, by POPs is a worldwide phe-nomenon (WHO, 1996) and measurable residues ofseveral organochlorine pesticides, PCBs, PCDDs andPCDFs are present in human tissues worldwide(Thomas and Colborn, 1992). Evidence for low-leveleffects of POPs on humans is more limited than that forwildlife but is consistent with effects reported both inexposed wildlife populations and in laboratory experi-ments on animals (UN-ECE, 1994).

7.3.1. Effects on the immune systemThe chronic effects of occupational, bystander and

near-field exposure to POP pesticides are difficult toassess. For example, few epidemiological studies havebeen carried out to assess changes in the human im-mune system from pesticide exposure. Moreover, al-though pesticide-induced immunosuppression mayincrease people’s susceptibility to infectious and para-sitic diseases, these diseases are so prevalent amonglow-income populations, and malnutrition and unsani-tary conditions are so widespread, that distinguishingany greater susceptibility due to immune deficiencies isdifficult (Repetto and Baliga, 1996). However, it hasbeen demonstrated that patients chronically exposed to

chlordane demonstrated clinical and immunologicalsymptoms highly suggestive of immune pathology andprobably a chlordane/heptachlor-induced autoimmunedisorder (Broughton et al., 1990). Nearly identical im-munological results were found in patients exposed tothe fungicide pentachlorophenol (McConnachie andZahalsky, 1991).

Immune modulating effects have also been detectedin people exposed to low-level, environmental concen-trations of POPs. Weisglas-Kuperus et al. (1995) foundcertain immunological aberrations were associated withpre and post-natal exposure of Dutch infants to PCBsand PCDDs. Although their data did not indicate thatthese aberrations caused any more illness among theinfants, they could persist and presage later difficultiessuch as immune suppression, allergy and autoimmunedisease. Swedish investigations have reported that di-etary intake of PCBs and PCDD/Fs may be linked toreductions in the population of natural killer cells(Svensson et al. (1993). These cells are believed to playa role in the body’s defence against viruses and tu-mours. However, none of the subjects in this studydisplayed any signs of health impairment attributableto lowered number of natural killer cells.

As mentioned earlier, POPs tend to migrate to higherlatitudes and to bioaccumulate and biomagnify, partic-ularly in marine ecosystems. Because fish, whale, seal,walrus and bear meat are mainstays of their diet, someCanadian Inuit consume relatively large amounts oforganochlorine pesticides, PCBs and PCDDs andbioaccumulate them in fatty tissues over years of expo-sure (Kinloch et al., 1992). The intake of PCBs andtoxaphene exceeds the ‘tolerable daily intake’ (TDI) formany Inuit consumers (Kinloch et al., 1992). In theearly 1990s the discovery of high levels of PCBs, DDTand HCB in Greenlanders’ adipose tissue (Table 4) andvery high levels of PCBs, DDE and mirex in Inuitbreast milk prompted concerns about their possibleadverse effects on human (particularly infant) health,including increased susceptibility to infections (De-wailly et al., 1993a; Mulvad et al., 1996). CanadianInuit babies who accumulated high doses oforganochlorines were significantly more likely to experi-ence acute otitis (infection of the middle ear) and wereharder to vaccinate since many failed to produce aprimary antibody response to the usual vaccines. Incomparison with bottle-fed babies, these breast milk-fed


babies demonstrated decreased ratios of helper-T cellsto effector-T cells that were correlated with duration ofbreast feeding and organochlorine levels in breast milk(Dewailly et al., 1993b). Infectious disease incidenceamong the Inuit children appeared to be associatedwith immune dysfunction (Birnbaum, 1995). Despitethis, Kinloch et al. (1992), in their study of the benefitsand risks of the Inuit diet, concluded that the knownbenefits of traditional foods and of breast feeding out-weigh the known adverse consequences of contaminantintake associated with the traditional diet. The WHOalso maintain that ‘where levels of some POPs in breastmilk approach or slightly exceed tolerable levels, breastfeeding should not be discouraged since the demon-strated significant benefits of this practice greatly out-weigh the small hypothetical risk that POPs may pose’(WHO, 1996). Although these recommendations aresupported in a recent AMAP (Arctic Monitoring andAssessment Programme) report, the authors also sug-gest that consideration should be given to developingdietary advice to promote the use of less-contaminatedtraditional food items which will also maintain nutri-tional benefits (AMAP, 1997).

7.3.2. Endocrine effectsSome researchers have suggested that falling human

sperm counts, the increasing incidence of certain repro-ductive abnormalities in the human male such as sper-matogenetic dysfunction, maldescent of testes(Garcıa-Rodrıguez et al., 1996), malformations of thepenis and testicular cancer, as well as breast cancer inwomen, may be linked to chronic exposure to low levelsof hormone-mimicking chemicals (including somePOPs). Toppari et al. (1996) have provided an extensivereview of the current knowledge about male reproduc-tive health and environmental xenoestrogens. However,at present there is an almost complete lack of exposuredata in humans to support such an association.

The incidence of breast cancer in women appears tobe rising in many countries and the pathogenesis ofbreast cancer has been linked with exposure to hor-mone disrupting chemicals (Davis et al., 1993). Falck etal. (1992) found higher levels of PCBs and DDE inmammary adipose tissue from women with malignantbreast cancer compared with women having benignbreast cancer. Wolff et al. (1993) demonstrated a posi-tive association between serum DDE (but not serumPCB) and incidence of breast cancer and Dewailly et al.(1994) observed higher levels of plasma PCB and DDEin women with estrogen receptor-positive breast cancer.On the other hand Krieger et al. (1994), in a muchlarger study, found neither DDE nor PCBs were associ-ated with breast cancer and Safe (1997) reported thatoccupational exposure to relatively high levels of PCBsand DDT/DDE was not associated with an increasedincidence of breast cancer. Key and Reeves (1994), after

producing a statistical summary of results from six suchstudies (including those mentioned above), concludedthat it was unlikely that DDT in the environmentincreased the risk of breast cancer and that there wasno evidence of such an association for PCBs.

7.3.3. De6elopmental, neural and beha6ioural effectsSigns of impaired neurological development in chil-

dren, through exposure to PCBs in utero and in theirmother’s milk, were demonstrated in a study of 866infants in North Carolina, USA (Rogan et al., 1986;Gladen et al., 1988). Further evidence that early expo-sure to some organochlorine pollutants can cause long-term intellectual impairment comes from a series ofstudies on a group of over 200 children, three-quartersof whom had mothers who had eaten significantamounts of fish known to be contaminated with PCBsand other pollutants from Lake Michigan. Differenceswere evident at birth with lower birth weight and asmaller head circumference the greater the mother’sconsumption of Lake Michigan fish (Jacobson et al.,1984). At 7 months of age there were signs of impairedcognitive function in those children exposed to higherPCB levels and at 4 years of age they achieved lowerscores in verbal and memory tests (Jacobson et al.,1990). At the age of 11 years these children were againassessed, this time using a range of tests for intelligencequotient (IQ), arithmetic, spelling, reading and compre-hension, and the results compared with their prenatalexposure to PCBs as measured in the placenta andmother’s blood and milk (Jacobson and Jacobson,1996). Their results show evidence of persistent intellec-tual impairment which has not been overcome by envi-ronment or education. They found that children fromthe group with highest PCB exposure had average IQlevels 6.2 points lower than children from the lessexposed groups as well as having poorer verbal compre-hension and being more easily distracted. Althoughmuch larger quantities of PCBs are transferred postna-tally through lactation than in utero, intellectual im-pairment occurred only in relation to transplacentalexposure. The authors concluded that these adverseeffects were due to PCB exposure in utero, one possiblemechanism being PCB-induced reduction in serum con-centrations of thyroid hormones which are needed tostimulate neuronal and glial proliferation and differen-tiation. They point out that the levels of PCBs carriedby the mothers were similar to or slightly above thegeneral population in the United States and thatwomen who do not eat fish may accumulate thesecompounds from a variety of other food sources.

In a longitudinal, prospective PCB/dioxin breast milkstudy, started in the Netherlands in 1990 and compris-ing 418 healthy mother/infant pairs, infants were exam-ined for endocrine, immune and neuro-behaviouraldevelopment. Higher PCDD, PCDF and PCB levels in


human milk (expressed as TEQs) correlated signifi-cantly with lower plasma levels of maternal total tri-iodothyronine and total thyroxine and higher plasmalevels of thyroid stimulating hormone (TSH) levels ininfants in the second week and third month after birth(Koopman-Esseboom et al., 1994). Infants exposed tohigher than average human milk TEQ levels also hadsignificantly lower plasma-free thyroxine and total thy-roxine levels, in the second week after birth, than thoseexposed to lower than average human milk TEQs. Itwas concluded that elevated levels of PCBs, PCDDsand PCDFs can alter the human thyroid status. Infantneuro-developmental testing showed a small negativeeffect of prenatal PCB exposure on the psychomotorscore at 3 months (Koopman-Esseboom et al., 1996).At 7 months of age both mental development andpsychomotor scores were positively correlated with du-ration of breast-feeding although for breast-fed infantsreceiving higher cumulative TEQs the positive effect ofbreast-feeding on psychomotor outcome was dimin-ished. Transplacental exposure to PCBs had a smallnegative effect on the neurological condition of 18-month old toddlers (Huisman et al., 1995).

8. Identifying additional POPs as candidates for futureinternational action

8.1. Introduction

Most man-made chemicals have been neither testednor evaluated for their hazard potential. There are100106 chemicals in the European inventory of existingchemical substances (EINECS) commonly referred toas ‘existing substances’. These are chemical substancesand entities which, by 1981, had been reported to andlisted by the EU-Commission as having been marketedduring the preceding 10 years. Of these, perhaps20000–30000 are currently marketed in significantquantities. ‘New substances’ (i.e. those substances notalready on the EINECS list) are being notified, but notnecessarily marketed, at a rate of about 200 a year andby the end of 1996 amounted to about another 2000substances (Bro-Rasmussen, 1996). The data and docu-mentation needed to evaluate many of their potentialeffects on human health and/or the environment existfor about 5000 chemicals. Only for a few hundred isthere sufficient knowledge for performing the full haz-ard/risk assessment described in the Commission of theEuropean Communities (1994) technical guidance doc-ument. Priority setting as a prelude to more refined orin-depth risk assessments is recognised as being ofcritical importance and the process used by the EU isdescribed by Van Leeuwen et al. (1996). The EU toxi-cological programme for ‘new substances’ involves dif-ferent levels of testing according to production volume

and use. Of the ‘existing chemical substances’ a starthas been made on providing data on the :2500HPVCs (High Production Volume Chemicals). Al-though data on most of the toxicity endpoints areavailable, only a very limited set of data is available forthe chronic endpoints described in the previous section.Effects on the endocrine system, for example, have onlybeen studied for a few of the man-made chemicalscurrently being used and it is likely that other estro-genic chemicals remain unidentified (Toppari et al.,1996). A much more focussed procedure, based ongroup prioritisation using criteria specific to POPs, willbe required.

8.2. Prioritization and assessment schemes

8.2.1. Proposed general risk assessment schemeConsidering the large number of untested chemical

substances that exist, of which POPs comprise a specificsubset requiring international joint action, it would beappropriate to establish a tiered selection and assess-ment procedure. A possible scheme for group priori-tization of existing chemicals is shown in Fig. 2. Undersuch a scheme, candidate POPs or PBTs could beplaced on a fast track for initial screening prior to amore costly and lengthy, full-scale risk assessment.Initial selection criteria for POPs would comprise long-range transport potential followed by persistence andbioaccumulation potential.

The risk assessment stage will then quantify theactual or predicted levels of environmental exposure tothe substance (exposure assessment) and the nature andlikely severity of any resulting adverse effects (effectsassessment). It is of the utmost importance that thetype of effects assessment used is appropriate for thegroup of chemicals in question. In the case of suspectedPOPs, this should include testing for the chronic effectsconsidered in Section 7, i.e. immunotoxicity, endocrinedisruption and carcinogenicity. When the levels of un-certainty involved in predicting exposure and the conse-quent possible effects on ecosystems, wildlife or humansare high, expert judgement and weight-of-evidence,based on the precautionary principle, may have to beused to decide whether or not the risks are acceptable.This would be especially so in the case of new sub-stances yet to be released into the environment forwhich future production, use and emission volumes areonly predictions and exposure assessments much moreuncertain.

8.2.2. Criteria proposed for the CLRTAP POPsprotocol

It is unlikely that the POPs chosen for initial inclu-sion in the CLRTAP protocol and the proposed UNEPglobal convention are the only persistent organic sub-stances released into the environment that have the


Fig. 2. Proposed general chemical risk assessment scheme.

potential to cause adverse effects in humans and/orthe environment at considerable distances from theirpoint of release. Crucial to the long-term success ofthe CLRTAP POPs protocol and the proposed UNEPglobal convention on POPs will be the criteria andprocedures chosen for adding new substances to theirrespective initial lists. Science-based criteria have beendeveloped for use within the CLRTAP POPs protocolto achieve this (Fig. 3).

If the risk profile is deemed acceptable by the execu-tive body, and further consideration of the substanceis determined to be warranted, one or more technicalreviews of the proposal will be conducted. These willbe based on the submission and any other relevantinformation submitted to the executive body. Expertjudgement will have to be used to determine, qualita-tively, the strength of the case for or against classify-ing the proposed substance as a protocol POP (IFCS,1996b).

8.2.3. Criteria to be used for the global UNEP POPscon6ention

Although the above criteria are likely to influence theprocedure to be used within the UNEP global conven-tion on POPs, the CLRTAP process is clearly onlyconcerned with long-range transboundary air pollution.The process used in the global convention is likely toincorporate a broader range of criteria including disper-sion in the hydrosphere and via migratory species ofanimals and the need to reflect possible influences ofmarine transport and tropical climates (Buccini, 1997).For example, triphenyl tin compounds were identifiedas high priority compounds after the first two stages ofthe CLRTAP selection process for the initial list ofprotocol POPs. However, after assessment in the thirdstage, these compounds were not included because theyare strongly associated with the aquatic phase and therewas no evidence of long-range atmospheric transport.Such compounds could be included under the global


Fig. 3. Information to be included in the risk profile of a substance a party proposes should be added to annexes I, II or III of the CLRTAP POPsprotocol. (Adapted from Executive Body decision 1998/2, UN-ECE, 1998b.)

POPs convention if the process set up by UNEP takeslong-range water transport into account. Clearly, thedefinition of a POP agreed for the CLRTAP protocolcould be modified to suit the purposes of a globalconvention.

8.2.4. Tests for chronic-toxicity endpointsGiven that a range of man-made chemicals (including

several POPs) have the capacity to disrupt the immuneand endocrine systems of both wildlife and humans, aswell as being potentially carcinogenic, any candidatePOP being considered within the proposed general riskassessment scheme (Fig. 2) or the CLRTAP procedure(Fig. 3) should be tested for these effects. Test proce-dures for several of the relevant endpoints for theseeffects are still being developed but are expected tobecome available within the next few years.

8.2.4.1. Immunotoxicity testing. Various tests exist fordetermining chemical immunotoxicity including in vitro(using tissue, cells or cell components outside the body)and in vivo (performed in living organisms) techniques.A two-tiered battery of tests has been recommended bythe US National Institute of Environmental HealthServices to gauge the potential for a chemical’s im-munotoxic effects on human health (Luster et al., 1988,1995; Repetto and Baliga, 1996). Tier 1 tests assesschanges in sensitive immune parameters. If these arefound, or if there is other evidence for immunotoxicity,Tier 2 tests are then performed in order to measurefunctional changes more fully (Table 5).

8.2.4.2. Testing for endocrine disruption. Although sev-eral existing regulatory tests are capable of detectingreproductive and developmental toxicants (ECETOC,


Table 5A two-tiered testing protocol for chemical immunotoxicity

Immune parameter Test

Immunopathology Pathology of immune organs (e.g. thymus, spleen)Tier 1Complete blood counts

Humoral immunity B lymphoproliferative responseAntibody levels

Cell-mediated immunity T lymphoproliferative responseNon-specific immunity Macrophage/neutrophil/natural killer cell activity

Tier 2 Immunopathology Differential blood counts (numbers and proportions of white blood cells and lymphocytes)Humoral immunity Secondary antibody responses (analogous to testing for effectiveness of vaccination against diseases)Cell-mediated immunity Delayed hypersensitivity (i.e. delayed onset of heightened responsiveness to an antigen following an

initial challenge)T cell cytosisPhagocytosis (by macrophages)Non-specific immunity

Host resistance models Bacterial challengeParasite challengeViral challengeTumour challenge

1996) they do not fully cover effects caused by en-docrine disrupters. The need to develop a hazard iden-tification strategy for endocrine disrupters wasaddressed at a recent workshop on ‘Endocrine Modula-tors and Wildlife: Assessment and Testing (EMWAT)’(Tattersfield et al., 1997). The hazard identificationstrategy they proposed involves three stages and utilisesa combination of structure activity relationship (SAR)mathematical models, in vitro tests and in vivo tests(Fig. 4).

In the future, SARs may become available for screen-ing and prioritizing possible endocrine-disrupting, im-munotoxic or carcinogenic chemicals for further, morecostly testing. SARs are based on the principle thatstructurally similar chemicals should have similar bio-logical activity and relate specifically-defined toxicologi-cal activity of chemicals to their molecular structureand physico–chemical properties. This requires a de-tailed in vitro toxicological knowledge base for theparticular mode of action in question. Given adequateresearch support, it is envisaged that SARs to predictthe binding affinities of the AhR, estrogen receptor(ER), androgen receptor (AR) and, possibly, thyroidreceptor should be available within three years (Tat-tersfield et al., 1997). However, models appropriate forother mechanisms that may have estrogenic effects invivo (e.g. changes in ligand synthesis) may take longer.The development of SARs for modes of action involv-ing the metabolism and transport of hormones are alsolikely in the future (Tattersfield et al., 1997). At theEMWAT workshop, it was concluded that high qualitySARs should be used as soon as they are availablerather than delay their use until all the mechanisms canbe modelled.

Several in vitro tests for endocrine disrupting poten-tial are available although they could be further refined

to enhance their reliability and selectivity. They arequick, relatively cheap to carry out, and can be used forelucidating mechanisms of action. In certain situations,they can also replace in vivo tests thus reducing theneed for animal testing. However, they are limited inpossessing only a portion of the metabolic systempresent in entire animals. Therefore, in the abovescheme, in vitro tests are seen as complimentary to, notsubstitutes for, in vivo tests. Before in vitro tests can beused routinely, suitable tests will have to be selected,optimised and validated using internationally agreedapproaches (Tattersfield et al., 1997).

In vivo tests have the advantage of being highlyintegrative and the capacity to evaluate mixed mecha-nisms of action and numerous endpoints. Short-term invivo tests can be rapid and inexpensive and are recom-mended for use in the screening/prioritizing stageabove, given that SARs and/or in vitro tests would beinadequate on their own. More lengthy, and hencemore costly, sub-chronic or chronic (multi-genera-tional) studies are recommended for use in the confir-mation stage. Although no currently available andvalidated in vivo tests are appropriate as they stand,several could be readily adapted, and others might formthe basis of useful tests after further development (Tat-tersfield et al., 1997). The possibility of unusual dose-re-sponse curves (e.g. inverted ‘U’) for endocrinedisrupters would have to be taken into account anddoses would, therefore, have to represent a full rangefrom high (possibly just sub-lethal) to low and includetypically five doses. Interspecies extrapolation wouldalso be problematic and any practical in vivo testingscheme, involving a very limited number of surrogatespecies, will inevitably produce false negative and falsepositive conclusions.


Because many of the test methods are still beingdeveloped, it is likely to be a few years yet before ascheme such as that described above becomes a usefuland reliable tool. The OECD (Organisation for Eco-nomic Co-operation and Development) EnvironmentalHealth and Safety Programme is co-ordinating activi-ties in this area in order to develop a common ap-proach to the identification and assessment ofendocrine disrupters. The newly established OECDWorking Group on Endocrine Disrupters Testing andAssessment (EDTA) met, in Paris in March 1998, toagree on objectives and priorities for international workon the development of methods for the testing andassessment of endocrine disrupters. A consensus on aconceptual framework for investigating chemicals forthese effects was reached, based on approaches recom-

mended at international workshops (such as that de-scribed above) and the US EDSTAC (EndocrineDisrupters Screening and Testing Advisory Commit-tee). Relevant endpoints to be included in the three tiersof the framework were agreed and the necessary techni-cal work on test methods referred to the expert groupsof the OECD Test Guidelines Programme. A validationproject was also launched, in which member countriesand international industry associations will co-operate,to be co-ordinated by OECD Management Teams con-sisting of balanced representation from member coun-tries’ governments, academia, industry and publicinterest groups.

8.2.4.3. Testing for carcinogenicity. This should com-prise in vitro tests on mutagenicity (microbial andmammalian mutagenicity tests, chromosome aberrationetc.) as well as intercellular communication and cellproliferation assays. When positive in either of the testsystems after repeated measurements, in vivo tumourpromotion and full carcinogenicity bioassays should beapplied.

9. Issues in assessing risk and selecting riskmanagement options for POPs

In schemes for identifying new POPs, such as thosedescribed in Figs. 2 and 3, it is necessary to draw adistinction between a scientific risk assessment andpolicy options relating to the subsequent managementof any risks so identified. Agreement on measures toreduce or eliminate these risks will only be achievable ifaccount is taken of the broader societal impact ofvarious measures. Considering that the benefits andrisks of substances, and the costs and benefits of mea-sures to manage or reduce those risks, are asymmetri-cally distributed, some form of comparison of costs andbenefits of various options will be needed. As is the casewith risk assessment, the level of this analysis must beadequate enough for informing the decision takers butnot so detailed as to delay the decisions indefinitely.

Deciding which management options are appropriateis only feasible if there is a good understanding of theefficacy and practicability of the measures (i.e. to whatextent will the measure actually reduce the risk), and onthe socio-economic impact of the proposed measure.Clearly, effective measures with minimal socio-eco-nomic impacts would be easy to agree upon whereas itmight be difficult to achieve agreement on measureswith insignificant effects and considerable socio–eco-nomic impacts.

While it would not be necessary to prescribe analyti-cal techniques, it would be useful to consider the rangeof possible measures and the different implementing

Fig. 4. Summary of hazard identification scheme for determination ofpotential endocrine modulating effects of new and existing substancesproposed at the EMWAT workshop, and reported by Tattersfield etal. (1997).


tools and the appropriate combination of tools andmeasures. Although these have been fairly well pro-vided for in the CLRTAP protocol, recommendedproduct restrictions on substances with considerableindustrial, agricultural or consumer application shouldbe supported by initiatives to make users aware of, andable to use, alternative, less hazardous substances ornon-chemical approaches. In any case, the suggestedalternative should also be compared with the originalusing a risk-benefit analysis to avoid the substitution ofa product we do not like by a product we do notunderstand.

10. Engaging the public and industry

It is clear from practical experience, as well as frommore theoretical considerations, that, to achieve fairand effective results in environmental policy it is cru-cially important to ensure effective involvement of dif-ferent stakeholders. This has been recognised within theUNEP POPs process and several NGOs, representingboth industry and public interest groups, have beenincluded in the initial IFCS ad hoc working group andsubsequent workshops.

The management and assessment of risks can beenhanced by broadening the knowledge base and dia-logue through the involvement of stakeholder commu-nities (including the public). The value of extending thepeer review community in this way has been docu-mented by several authors (Funtowcz and Ravetz,1991; Renn, 1995; Wynne, 1996) and a range of meth-ods for facilitating participation have been established(Renn et al., 1995). For example, integrated environ-mental assessment (IEA) is an approach that allows abroader participation in environmental assessment anddecision-making. IEA offers a method of constructingand using scientific knowledge in environmental policythat can not only be peer reviewed scientifically but canalso be scrutinised by the stakeholders (e.g. publicadministrators, community and non-governmental rep-resentatives, business people and the media) during theassessment process, not just afterwards (Bailey, 1997).In the area of POPs control, some form of IEA mightprove to be a useful aid for decision-makers, in con-junction with more conventional risk assessments.

Perception of environmental and ecological risks hasnot been extensively studied outside the human healthcontext (Sjoberg, 1995). The differences between expertand public perceptions of the scale of the POPs riskscould be manifested in different ways (as described bySjoberg). A lack of trust in experts regarding POPscontrol, coupled with a fear of chemicals in traceamounts that one cannot see, smell or taste, might beexpected to result in a situation whereby risks judged tobe very small by experts are believed by lay people to

be quite large. On the other hand, the lack of overtdamage to human health (in most populations) coupledwith a general lack of public knowledge about POPs,may result in the opposite situation with the greaterrisk being perceived by experts. In the case of POPs,differences between other interest groups, such as envi-ronmental pressure groups versus industry representa-tives, may be of greater relevance. Alternatively, theremay be significant differences in perception betweenpopulations in different regions depending on the ex-tent to which they believe they benefit from, or areharmed by, POPs (e.g. people living in areas whereDDT is used successfully against malaria versus certainInuit groups in Canada exposed to PCB originatingfrom remote sources).

Improvements in the availability of information re-lating to products would support this process. Manycompanies have joined ‘Responsible Care’; a voluntarychemical industry organization devoted to environmen-tally sound production methods and life-cycle productstewardship. However, there maybe scope for furthervoluntary activity by the chemical and related indus-tries including support for health research programmesand sharing results of laboratory toxicity tests. Manycompanies (mostly outside the chemical industry), haveadopted the principles of ‘The Natural Step’, an inter-national movement dedicated to helping society reduceits impact on the environment and move towards asustainable future. It began in Sweden in 1989, whereover 25 of the largest corporations have now used TheNatural Step training to modify operations in accor-dance with four ‘system conditions for sustainability’.Companies in other European countries, including TheNetherlands and the UK, are increasingly adopting theprinciples of this approach as are many in the USA.The system conditions of The Natural Step are:1. Materials from the earth’s crust must not systemati-

cally increase in nature.2. Persistent substances produced by society must not

systematically increase in nature.3. The physical basis for the earth’s productive natural

cycles and biological diversity must not be systemat-ically deteriorated.

4. There must be fair and efficient use of resourceswith respect to meeting human needs.

Conditions 2 and 3 are of particular relevance for thecontrol of POPs and are in line with the objectives ofthe Esbjerg Declaration. Voluntary initiatives such as‘Responsible Care’ and ‘The Natural Step’ should beseen as complementary to the regulatory approachesfor controlling POPs and other hazardous substances.They also offer a means of broadening participation toinclude enlightened companies and, ultimately, an in-formed public which can, through their purchasingchoices, encourage the wider adoption of theseprinciples.


11. Recommendations

Data(1) For both model validation and risk assessment

purposes, reliable source and emissions data are re-quired, not only for existing POPs, but also for PBTsand hazardous substances in general.

(2) Procedures for gaining such data need to beharmonised and co-ordinated at an international level.

(3) Reliable monitoring data are needed for riskassessment and model calibrations. Such data fromdifferent regions should be comparable.

Models(4) There is a need to improve and validate models

of chemical fate in the environment at the local, re-gional, continental and global scale so that exposurescan be estimated with more reliability, especially for‘new’ chemicals and those for which inadequate moni-toring data exist.

Processes and effects(5) Further research is required on deposition/emis-

sion processes, transformation processes and bioavail-ablity of POPs and PBTs in terrestrial ecosystems.

(6) Further research is required on the phytotoxiceffects of POPs and PBTs and their effects on soilmicrobial populations and soil fauna.

(7) Further research is needed on transport pro-cesses, persistence, fate and effects of POPs and PBTsin aquatic media.

Effects (general)(8) Research is required into the effects of mixtures

of chemicals similar to those actually encountered inthe environment.

(9) Efforts should be increased to develop mecha-nism-based biomarkers of effect, as well as biologicallybased sum-parameters to aid the development of rele-vant human exposure and effects monitoring data.

(10) Further studies are recommended on effectmonitoring for POPs making use of existing acciden-tally high exposed populations (e.g. victims of theSeveso accident) in comparison with those exposed tothe highest background levels (e.g. the Inuit).

(11) Reliable, internationally agreed testingmethodologies for endocrine disrupters identificationand for immunotoxicity should be developed.

Chemical properties(12) International activities aimed at determining

and correlating the key physical, chemical and biologi-cal properties of chemicals used for assessment pur-poses need to be improved and better co-ordinated.

(13) There is a need for accepted test methods forlong-term persistence as applied to POPs and PBTs.

Monitoring(14) A systematic international search of environ-

mental media for the presence of ‘new’ and ‘unex-pected’ organic contaminants is required. (To date,

there has been undue reliance on fortuitous detection ofthese substances.)

Risk management(15) When it is judged that a chemical warrants

risk management, the actions required should be clearlydefined and a schedule of actions, and the correspond-ing expected beneficial reductions in concentrations andexposure, should be documented. Progress towardsthese goals should be reported regularly.

Alternati6es to POPs(16) As a priority, the availability of safer alterna-

tive chemicals or alternative practices, as well as train-ing in their use, will have to be facilitated, especially indeveloping countries. A comprehensive risk-benefitanalysis should be the basis for the development ofsuch alternatives.

Co-operation(17) There is a need for close co-operation between

the various international initiatives on POPs to ensureefficient use of scarce resources by, for example, ensur-ing that assessment criteria and procedures are as com-patible as possible.

Acknowledgements

This paper is the outcome of a workshop held fromMay 5–8, 1998, in York, UK under the auspices of theEuropean Training and Assessment Foundation(ETAF) in collaboration with the Stockholm Environ-ment Institute at York (SEI-Y). The authors would liketo acknowledge the valuable contributions made byProfessor Gordon Goodman, Nicole Wevers and theobservers at the workshop, Agneta Sunden-Bylehn(United Nations Environmental Programme (UNEP)Chemicals, Geneva) and Nathalie Scheidegger (Min-istry of Agriculture, Nature Management and Fisheries,The Netherlands).

References

Aguilar, A., Raga, J.A., 1993. The striped dolphin epizootic in theMediterranean Sea. Ambio 22, 524–528.

Alba, O.M., 1988. A preliminary study on exposure of rice farmers topesticides. In: Teng, P.S., Heong, K.L. (Eds.), Pesticides Manage-ment and Integrated Pest Management in Southeast Asia. Con-sortium for International Crop Protection, College Park,Maryland, pp. 417–421.

Alsberg, T., Balk, L., Nylund, K., de Wit, C., Bignert, A., Olsson,M., Odsjo, T., 1993. Persistent organic pollutants and the envi-ronment, SNV Report 4246, Swedish Environmental ProtectionAgency.


AMAP, 1997. Arctic pollution issues: a state of the Arctic environ-ment report Arctic Monitoring and Assessment Programme, Oslo,Norway.

Atlas, E., 1990. The long-range transport of organic compounds. In:Knap, A.H. (Ed.), The Long-range Atmospheric Transport ofNatural and Contaminant Substances. Kluwer, Dordrecht, pp.105–135.

Aulerich, R., Ringer, R., 1977. Current status of PC toxicity to mink,and after-effects on their reproduction. Arch. Environ. Contam.Toxicol. 6, 279–292.

Baek, S.O., Field, R.A., Goldstone, M.E., Kirk, P.W., Lester, J.N.,Perry, R., 1991. A review of polycyclic aromatic hydrocarbons-sources, fate and behaviour. Water, Air and Soil Pollution 60,279.

Bailey, P.D., 1997. IEA: a new methodology for environmentalpolicy? Environ. Impact Assess. Rev. 17, 221–226.

Balash, K.J., Al-Omar, M.A., Latif, B.M.A., 1987. Effect of chlor-dane on testicular tissues of Swiss mice. Bull. Environ. Contam.Toxicol. 39, 434–442.

Ballschmitter, K., 1996. Persistent, ecotoxic, and bioaccumulatingcompounds and their possible environmental effects. Pure andAppl. Chem. 68, 1771–1780.

Banerjee, B., 1987. Effects of sub-chronic DDT exposure on humoraland cell-mediated immune responses in albino rats. Bull. Environ.Contam. Toxicol. 39, 827–834.

Banerjee, B., Ramachandran, M., Hussain, Q., 1986. Subchroniceffect of DDT on humoral immune response in mice. Bull.Environ. Contam. Toxicol. 37, 433–440.

Barnett, J.B., Rodgers, K.E., 1994. Pesticides. In: Dean, J.H., Luster,M.I., Munson, A.E., Kimber, I. (Eds.), Immunotoxicology andImmunopharmacology, 2nd Edition. Raven Press, NY, pp. 191–211.

Barnett, J.B., Barfield, L., Walls, R., Joyner, R., Owens, R., Soder-burg, L.S.F., 1987. The effects of in utero exposure to hex-achlorobenzene on the developing immune response of BALB/cmice. Toxocol. Lett. 39, 263–274.

Barrie, L.A., 1986. Arctic air pollution: an overview of currentknowledge. Atmos. Environ. 20, 643–663.

Barrie, L.A., Gregor, D., Hargrave, B., Lake, R., Muir, D., Shearer,R., Tracey, B., Bidleman, T., 1992. Arctic contaminants: sources,occurrence and pathways. Sci. Tot. Env. 122, 1–74.

Berdowski, J.J.M., Bass, J., Bloos, J.P.J., Visschedijk, A.J.H., Zand-veld, P.Y.J., 1997. The European emission inventory of heavymetals and persistent organic pollutants, TNO report UBA-FB,UFOPLAN Ref. no. 104.02 672/03, Apeldoorn, p. 239.

Bergman, A., Klasson-Wehler, E., Kuroki, H., 1994. Selective reten-tion of hydroxylated PCB metabolites in blood. Environ. HealthPerspec. 102, 464–469.

Bidleman, T.F., Billings, W.N., Foreman, W.T., 1986. Vapor-particlepartitioning of semi-volatile organic compounds: Estimates fromfield collections. Environ. Sci. Technol. 20, 1038–1043.

Bidleman, T.F., Patton, G.W., Walla, M., Hargrave, B., Vass, W.,Erickson, P., Fowler, B., Scott, V., Gregor, D.J., 1989. Toxapheneand other organochlorines in Arctic ocean fauna: evidence foratmospheric delivery. Arctic 42, 307–313.

Bidleman, T.F., Walla, M.D., Roura, R., Carr, E., Schmidt, S., 1993.Organochlorine pesticides in the atmosphere of the SouthernOcean and Antarctica, January-March, 1990. Mar. Poll. Bull. 26,258–262.

Bidleman, T.F., Jantunen, L.M., Falconer, R.L., Barrie, L.A., 1995.Decline of hexachlorocyclohexane in the arctic atmosphere andreversal of air-sea gas exchange. Geophys. Res. Lett. 22, 219–222.

Birnbaum, L.S., 1995. Workshop on perinatal exposure to dioxin-likecompounds: V. immunologic effects. Environ. Health Perspec.103, 157–159.

Borrell, A., Aguilar, A., 1991. Were PCB levels in striped dolphinsaffected by the western Mediterranean die-off abnormally high?Eur. Res. Cetaceans 5, 88–92.

Bowes, G.W., Jonkel, C.J., 1975. Presence and distribution of poly-chlorinated biphenyls (PCB) in Arctic and subarctic marine foodchains. J. Fish. Res. Board Can. 32, 2111–2123.

Brandt, I., Bergman, A., 1987. PCB methyl sulphones and relatedcompounds: Identification of target cells and tissues in differentspecies. Chemosphere 16, 1671–1676.

Brandt, I., Lund, J., Bergman, A., Klasson-Wehler, E., Poellinger L.,Gustafsson, J.-A., 1985. Target cells for the PCB metabolite4,4%-bis(methylsulphonyl)- 2,2%,5,5%-tetrachlorobiphenyl in lungand kidney. Drug Metabolism Disp. 13, 490–496.

Brandt, I., Jonsson, C.-J., Lund, B.-O., 1992. Comparative studies onadrenocorticolytic DDT-metabolites. Ambio 21, 602–605.

Bro-Rasmussen, F., 1996. Chemical hazard is a serious matter-tooserious to be left for chemists to administer, (Translated fromDanish) Aktuelt Miljo, No. 6, (J.H. Schultz Information Ltd.).

Broughton, A., Thrasher, J.D., Madison, R., 1990. Chronic healtheffects and immunological alterations associated with exposure topesticides. Comments Toxicol. 4, 59–71.

Brouwer, A., Van den Berg, K.J., 1986. Binding of a metabolite of3,4,3%,4%-tetrachlorobiphenyl to transthyretin reduces serum vita-min-a transport by inhibiting the formation of the protein com-plex carrying both retinol and thyroxine. Toxicol. Appl.Pharmacol. 85, 301–312.

Brouwer, A., Reijnders, P.J.H., Koeman, J.H., 1989. Polychlorinatedbiphenyl (PCB)-contaminated fish induce vitamin A and thyroidhormone deficiency in the common seal (Phoca 6itulina). Aquat.Toxicol. 15, 99–106.

Brouwer, A., Ahlborg, U.G., Van den Berg, M., Birnbaum, L.S.,Boersma, E.R., Bosveld, B., Denison, M.S., Gray, L.E., Hagmar,L., Holene, E., Huisman, M., Jacobson, S.W., Jacobson, J.L.,Koopman-Esseboom, C., Koppe, J.G., Kulig, B.M., Morse, D.C.,Muckle, G., Peterson, R.E., Sauer, P.J.J., Seegal, R.F., Smits vanProoije, A.E., Touwen, B.C.L., Weisglas-Kuperus, N., Winneke,G., 1995. Functional-aspects of developmental toxicity of poly-halogenated aromatic-hydrocarbons in experimental animals andhuman infants. Eur. J. Pharmacol. (Environ. Toxicol. section)293, 1–40.

Brouwer, A., Morse, D.C., Lans, M.C., Schuur, A.G., Murk, A.J.,Klasson-Wehler, E., Bergman, A., Visser, T.J., 1998. Interactionsof persistent environmental organohalogens with the thyroid hor-mone system: Mechanisms and possible consequences for animaland human health. Toxicol. Ind. Health 14, 59–84.

Buccini, J., 1997. Persistent organic pollutants (POPs): Recent devel-opments in the Intergovernmental Forum for Chemical Safety(IFCS), UNEP, Geneva. (http://irptc.unep.ch/pops/stpeter/stpeter1.html)

Calamari, D., Bacci, E., Focardi, S., Gaggi, C., Morosini, M., Vighi,M., 1991. Role of plant biomass in the global environmentalpartitioning of chlorinated hydrocarbons. Environ. Sci. Technol.25, 1489–1495.

Clausen, J., Berg, O., 1975. The content of polychlorinated hydrocar-bons in Arctic ecosystems. Pure Appl. Chem. 42, 223–232.

Colborn, T., 1991. Epidemiology of Great Lakes bald eagles. J.Toxocol. Environ. Health 33, 395–453.

Commission of the European Communities, 1994. Technical guidancedocuments in support of the risk assessment regulation (1488/94)for existing substances in the context of Council Regulation793/93/EEC, Commission of the European Communities, Brus-sels, Belgium.

Cowan, C.E., Mackay, D., Feijtel, T.C.J., van de Meent, D., DiGuardo, A., Davies, J., Mackay, N., 1995. The multi-media fatemodel: a vital tool for predicting the fate of chemicals. Proceed-ings of the Society of Environmental Toxicology and Chemistry(SETAC) workshop held in Leuven, Belgium, April 14-16, 1994and Denver, Colorado, November 4-5, 1994. SETAC, Pensacola,FL.

Daisey, J.M., McCaffrey, R.J., Gallagher, R.A., 1981. Polycyclicaromatic hydrocarbons and extractable particulate organic matterin the Arctic aerosol. Atmos. Environ. 15, 1353–1364.


Darnerud, P.O., Morse, D.C., Klasson-Wehler, E., Brouwer, A.,1996. Binding of a 3,3%,4,4%-tetrachlorobiphenyl (CB–77) metabo-lite to fetal transthyretin and effects on fetal thyroid hormonelevels in mice. Toxicology 106, 105–114.

Davis, D., Safe, S.H., 1989. Dose-response immunotoxicities of com-mercial polychlorinated biphenyls (PCBs) and their interactionwith 2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol. Lett. 48, 35–43.

Davis, D.L., Bradlov, H.L., Wolff, M., Woodruff, T., Hoel, D.G.,Anton-Culver, H., 1993. Medical hypothesis: xenoestrogens aspreventable causes of breast cancer. Environ. Health Perspect.101, 372–377.

De Guise, S., Lagace, A., Beland, P., 1994. Tumours in St. LawrenceBeluga Whales (Delphinapterus leucas). Veterinary Path. 31, 444–449.

De Guise, S., Martineau, D., Beland, P., Fournier, M., 1995. Possiblemechanisms of action of environmental contaminants on St.Lawrence Beluga Whales (Delphinapterus leucas). Environ. HealthPerspect. 103, 73–77.

De Swart, R.L., Kluten, R.M.J., Huizing, C.J., Vedder, L.J., Reijn-ders, P.J.H., Visser, I.K.G., Uytde-Haag, F.G.C.M., Osterhaus,A.D.M.E, 1993. Mitogen and antigen induced B and T cellresponses of peripheral blood mononuclear cells from the harbourseals (Phoca 6itulina). Veterinary Immunol. Immunopath. 37,217–230.

De Swart, R.L., Ross, P.S., Vedder, L.J., Timmerman, H.H., Heis-terkamp, S.H., Van Loveren, H., Vos, J.G., Reijnders, P.J.H.,Osterhaus, A.D.M.E, 1994. Impairment of immune function inharbour seals (Phoca 6itulina) feeding on fish from pollutedwaters. Ambio 23, 155–159.

De Swart, R.L., Ross, P.S., Timmerman, H.H., Vos, H.W., Reijn-ders, P.J.H., Vos, J.G., Osterhaus, A.D.M.E, 1995. Impairedcellular immune response in harbour seals (Phoca 6itulina) feedingon environmentally contaminated herring. Clin. Exp. Immunol.101, 480–486.

Dewailly, E., Nantel, A., Weber, J.P., Meyer, F., 1989. High levels ofPCBs in breast milk of Inuit women from Arctic Quebec. Bull.Environ. Contam. Toxicol. 43, 641–646.

Dewailly, E., Ayotte, P., Bruneau, S., Laliberte, C., Muir, D.C.G.,Norstrom, R.J., 1993a. Inuit exposure to organochlorines throughthe aquatic food chain in arctic Quebec. Environ. Health Perspec.101, 618–620.

Dewailly, E., Bruneau, S., Laliberte, C., Belles-Iles, M., Weber, J.P.,Ayotte, P., Roy, R., 1993b. Breast milk contamination by PCBsand PCDDs/PCDFs in arctic Quebec: preliminary results on theimmune status of Inuit infants. in: H. Fielder, H. Frank, O.Hutzinger, W. Parzefall, A. Riss and S. Safe (Eds.), Dioxin ’93:13th International Symposium on Chlorinated Dioxins and Re-lated Compounds. Federal Environmental Agency, Vienna, Sep-tember 1993 Organohalogen Compounds, Vol. 13, pp. 403–406.

Dewailly, E., Dodin, S., Verreault, R., Ayotte, P., Sauve, L., Morin,J., Brisson, J., 1994. High organochlorine body burden in womenwith estrogen receptor-positive breast cancer. J. Natl. Cancer Inst.86, 232–234.

Dietz, R., Heide-Jørgensen, M.P., Harkonen, T., 1989. Mass deathsof harbour seals (Phoca 6itulina) in Europe. Ambio 18, 258–264.

Duursma, E.K., Carroll, J., 1996. Environmental compartments.Equilibria and Assessment of Processes Between Air, Water,Sediments and Biota. Springer, Germany.

ECETOC (European Centre for Ecotoxicology and Toxicology ofChemicals) 1996, Environmental Oestrogens Task Force–Com-pendium of test methods, Document No. 33, Brussels, Belgium.

Eriksson, G., Jensen, S., Kylin, H., Strachan, W., 1989. The pineneedle as a monitor of atmospheric pollution. Nature 341, 42–44.

Eriksson, P., 1992. Neuroreceptor and behavioural effects of DDTand pyrethroids in immature and adult mammals. In: Isaacson,

R.L., Jensen, K.F. (Eds.), The Vulnerable Brain and Environmen-tal Risks, vol. 2. Plenum Press, NY, pp. 235–251.

Eriksson, P., 1997. Developmental neurotoxicity of environmentalagents in the neonate. Neurotoxicology 18, 719–726.

Eriksson, P., Fredriksson, A., 1996. Developmental neurotoxicity offour ortho-substituted polychlorinated biphenyls in the neonatalmouse. Environ. Toxicol. Pharmacol. 1, 155–165.

European Commission, 1996. European workshop on the impact ofendocrine modulators on human health and wildlife, Weybridge,UK, Report of the proceedings, EUR 17549.

Facemire, C., Gross, T., Guillette, L., 1995. Reproductive impairmentin the Florida Panther: nature or nurture. Environ. Health Per-spect. Suppl. 103, 79–86.

Falck Jr., F., Ricci Jr., A., Wolff, M.S., Godbold, J., Deckers, P.,1992. Pesticides and polychlorinated biphenyl residues in humanbreast lipids and their relation to breast cancer. Arch. Environ.Health 47, 143–146.

Fiedler, H., 1997. Polychlorinated biphenyls (PCB), Presentation tothe UNEP sub-regional meeting on identification and assessmentof releases of persistent organic pollutants, St. Petersburg, Rus-sian Federation, 1–4 July 1997. UNEP, Geneva. (http://irptc.unep.ch/pops/stpeter/stpete2c.html)

Finizio, A., Bidleman, T.F., Szeto, S.Y., 1997a. Emission of chiralpesticides from an agricultural soil in the Fraser Valley, BritishColumbia. Chemosphere 36, 345–355.

Finizio, A., Mackay, D., Bidleman, T., Harner, T., 1997b. Octanol-air partition coefficient as a predictor of partitioning of semi-volatile organic chemicals to aerosols. Atmos. Environ. 31,2289–2296.

Fox, G.A., 1992. Epidemiological and pathobiological evidence ofcontaminant-induced alterations in sexual development in free-liv-ing wildlife. In: Colborn, T., Clement, C. (Eds.), Chemically-In-duced Alterations in Sexual and Functional Development: TheWildlife/Human Connection. Princeton Scientific Publishing,Princeton, NJ, pp. 147–158.

Fry, D.M., Toone, C.K., 1981. DDT-induced feminization of gullembryos. Science 213, 922–924.

Funtowicz, S.O., Ravets, J.R., 1991. A new scientific methodologyfor global environmental issues. In: Costanza, R. (Ed.), EcologicalEconomics: The Science and Management of Sustainability. Co-lumbia University Press, New York, pp. 137–152.

Gaebler, O.H., Vitti, T.G., Vumirovich, T., 1966. Isotope effects inmetabolism of 15N and 14 N from unlabelled dietary proteins. J.Biochem. Phys. 44, 1245–1257.

Gagosian, R.B., Peltzer, E.T., Zafiriou, O.C., 1981. Atmospherictransport of continentally derived lipids to the tropical NorthPacific. Nature 291, 312–314.

Garcıa-Rodrıguez, I.R., Garcıa-Martın, M., Nogueras-Ocana, M., deDios Luna-del-Castillo, J., Espigares Garcıa, M., Olea, N.,Lardella-Claret, P, 1996. Exposure to pesticides and cryp-torchidism: geographical evidence of a possible association. Envi-ron. Health Perspect. 104, 1090–1095.

George, J.L., Frear, D.E.H., 1966. Pesticides in the Antarctic. J.Appl. Ecol. 3, 155–167.

Gilbertson, M., Kubiak, T., Ludvig, J., Fox, G., 1991. Great Lakesembryo mortality, edema and deformities syndrome (GLEMEDS)in colonial fish-eating waterbirds: similarity to chick edema dis-ease. J. Toxicol. Environ. Health 33, 455–520.

Gladen, B., Rogan, W., Hardy, P., Thullen, J., Tingelstad, J., Tully,M., 1988. Development after exposure to polychlorinatedbiphenyls and dichlorodiphenyl dichloroethene transplacentallyand through human milk. J. Pediatr. 113, 991–995.

Goldberg, E.D., 1975. Synthetic organohalides in the sea. Proc.Royal Soc. Lond. 189, 277–289.

Golden, R.J., Noller, K.L., Titus-Ernstoff, L., Kaufman, R.H., Mit-tendorf, R., Stillman, R., Reese, E.A., 1998. Environmental en-docrine modulators and human health: An assessment of thebiological evidence. Crit. Rev. Toxicol. 28, 109–227.


Government of Canada, 1991. Toxic chemicals in the Great Lakesand associated effects, Minister of supply and services, CatalogueNumber En 37-95/1990-1E.

Grachev, M.A., Kumarev, V.P., Mamaev, L.V., Zorin, V.L., Bara-nova, L.V., Denikina, N.N., Belikov, S.I., Petro, E.A., Kolesnik,V.S., Kolesnik, R.S., Dorofeev, V.M., Beim, A.M., Kudelin,V.N., Nagieva, F.G., Siderov, V.N., 1989. Distemper virus inBaikal seals. Nature Lond. 338, 209.

Grasman, K.A., 1995. Developmental immunotoxicity of environ-mental contaminants in fish-eating birds of the Great Lakes,Abstract presented at the Conference, Chemically-induced Alter-ations in the Developing Immune System: The Wildlife/HumanConnection, Racine, Wisconsin, February 10–12, 1995.

Gray Jnr., L.E., 1992. Chemical-induced alterations of sexual differ-entiation: a review of effects in humans and rodents. In: Colborn,T., Clement, C. (Eds.), Chemically-Induced Alterations in Sexualand Functional Development: The Wildlife/Human Connection.Princeton Scientific Publishing, Princeton, NJ, pp. 203–230.

Gregor, D.J., Gummer, W.D., 1989. Evidence of atmospheric trans-port and deposition of organochlorine pesticides and polychlori-nated biphenyls in Canadian Arctic snow. Environ. Sci. Technol.23, 561–565.

Guillette, L.J., Gross, T.S., Masson, G.R., Matter, J.M., Percival,H.F., Woodword, H.R., 1994. Developmental abnormalities ofthe gonad and abnormal sex hormone concentrations in juvenilealligators from contaminated and control lakes in Florida. Envi-ron. Health Perspect. 102, 680–688.

Han, S-L., Stone, D., 1997. Effects overview of persistent organicpollutants in the context of effects-based approaches to manage-ment, Background document for: The Workshop on CriticalLimits and Effects Based Approaches for Heavy Metals andPersistent Organic Pollutants, Bad Harzburg, Germany, Novem-ber 3-7, UN-ECE Convention on Long-range Transboundary AirPollution, Task Force on Mapping.

Hargrave, B.T., Vass, W.P., Erickson, P.E., Fowler, B.R., 1988.Atmospheric transport of organochlorines to the Arctic Ocean.Tellus 40B, 480–493.

Henny, C. and R. Grove, 1996. Hypoplastic reproductive organsrelated to xenobiotic compounds in young male River Otter fromthe Columbia River. in: Proceedings of the 17th annual meeting ofthe Society of Environmental Toxicology and Chemistry. Wash-ington DC, USA, November 1996.

Hoff, R.M., Muir, D.C.G., Grift, N.P., 1992. Annual cycle of poly-chlorinated biphenyls and organohalogen pesticides in air inSouthern Ontario. 2. Atmospheric transport and sources. Envi-ron. Sci. Technol. 26, 276–283.

Holsapple, M.P., Snyder, N.K., Wood, S.C., Morris, D.L., 1991. Areview of 2,3,7,8-tetrachlorodibenzo-p-dioxin- induced changes inimmunocompetence: 1991 update. Toxicology 69, 219–255.

Hornbuckle, K.C., Eisenreich, S.J., 1996. Dynamics of gaseoussemivolatile organic compounds in a terrestrial ecosystem – ef-fects of diurnal and seasonal climate variations. Atmos. Environ.30, 3935–3945.

Horstmann, M., Bopp, U., McLachlan, M.S., 1997. Comparison ofthe bulk deposition of PCDD/F in a spruce forest and an adjacentclearing. Chemosphere 34, 1245–1254.

Howsham, M., Jones, K.C., 1998. Sources of PAHs in the environ-ment. In: Nielsen, A.H. (Ed.), Handbook of EnvironmentalChemistry, Vol 3, Part 1, PAHs and Related Compounds.Springer-Verlag, Berlin, Heidlberg, pp. 137–174.

Huisman, M., Koopman, C., Lanting, C.I., Vanderpaauw, C.G.,Tuinstra, L.G.M.T., Fidler, V., Weisglas-Kuperus, N., Sauer,P.J.J., Boersma, E.R., Touwen, B.C.L., 1995. Neurological condi-tion in 18-month-old children perinatally exposed to polychlori-nated-biphenyls and dioxins. Early Hum. Develop. 43, 165–176.

IFCS, 1996a. Persistent organic pollutants: Socioeconomic consider-ations for global action – Theme paper prepared for an IFCS

expert meeting on persistent organic pollutants, Manila, thePhilippines, 17–19 June 1996, Report no. IFCS/EXP.POPs.2(UNEP, Geneva). (http://irptc.unep.ch/pops/indxhtms/manexp3.html)

IFCS, 1996b. Discussion paper on the development of science-basedcriteria for identifying further POPs, United Kingdom, Paper forIFCS meeting, Manila, 20–21 June 1996, Intergovernmental Fo-rum on Chemical Safety, UNEP (c/o WHO) Geneva. (http://irptc.unep.ch/pops/indxhtms/manwg5.html)

Iwata, H., Tanabe, S., Sakai, N., Tatsukawa, R., 1993. Distributionof persistent organochlorines in the oceanic air and surface seawa-ter and the role of ocean on their global transport and fate.Environ. Sci. Technol. 27, 1080–1098.

Iwata, H., Tanabe, S., Sakai, N., Nishimura, A., Tatsukawa, R.,1994. Geographical distribution of persistent organochlorines inair, water and sediments from Asia and Oceania, and theirimplications for global redistribution from lower latitudes. Envi-ron. Pollut. 85, 15–33.

Jacobson, J.L., Jacobson, S.W., 1996. Intellectual impairment inchildren exposed to polychlorinated biphenyls in utero. New Engl.J. Med. 335, 783–789.

Jacobson, J., Jacobson, S., Schwartz, P., Fein, G., Dowler, J., 1984.Prenatal exposure to an environmental toxin: a test of the multi-ple effects model. Developmental Psychol. 20, 523–532.

Jacobson, J.L., Jacobson, S.W., Humphrey, H.E.B., 1990. Effects ofin utero exposure to polychlorinated biphenyls and related con-taminants on cognitive functioning in young children. J. Pediatr.116, 38–45.

Jansen, H.T., Cooke, P.S., Porcelli, J., Liu, T.C., Hansen, L.G., 1993.Estrogenic and antiestrogenic actions of PCBs in the female rat –in vitro and in vivo studies. Reprod. Toxicol. 7, 237–248.

Jensen, S., Jansson, B., 1976. Anthropogenic substances in seal fromthe Baltic: Methyl sulphone metabolites of PCB and DDT. Am-bio 5, 257–260.

Jensen, S., Kihlstrom, J.E., Olsson, M., O8 rberg, J., 1977. Effects ofPCB and DDT on mink (Mustela 6ision) during the reproductiveseason. Ambio 6, 239.

Jonsson, C.-J., Rodriguez-Martinez, H., Lund, B.-O., Bergman, A.,Brandt, I., 1991. Adrenocortical toxicity of 3-methylsulfonyl-DDE in mice. II. Mitochondrial changes following ecologicallyrelevant doses. Fundam. Appl. Toxicol. 16, 365–374.

Jonsson, C.J., Lund, B.-O., Bergman, A., Brandt, I., 1992. Adreno-cortical toxicity of 3-methylsulhonyl-DDE (III): Studies in fetaland suckling mice. Reproductive Toxicol. 6, 233–240.

Kennedy, S., Smyth, J.A., McCullough, S.J., Allan, G.M., McNeilly,F., McQuaid, S., 1988. Confirmation of cause recent seal deaths.Nature Lond. 335, 404.

Key, T., Reeves, G., 1994. Organochlorines in the environment andbreast cancer. Brit. Med. J. 308, 1520–1521.

Kidd, K.A., Schindler, D.W., Hesslein, R.H., Muir, D.C.G., 1995.Correlation between stable nitrogen isotope ratios and concentra-tions of organochlorines in biota from a freshwater food web. Sci.Tot. Environ. 160/161, 381–390.

Kinloch, D., Kuhnlein, H., Muir, D.C.G., 1992. Inuit foods and diet:a preliminary assessment of benefits and risks. Sci. Tot. Environ.122, 247–278.

Koopman-Esseboom, C., Morse, D.C., Weisglas-Kuperus, N.,Lutkeschipholt, I.J., Van der Paauw, C.G., Tuinstra, L.G.M.T.,Brouwer, A., Sauer, P.J.J., 1994. Effects of dioxins and polychlo-rinated-biphenyls on thyroid-hormone status of pregnant-womenand their infants. Pediatr. Res. 36, 468–473.

Koopman-Esseboom, C., Weisglas-Kuperus, N., Deridder, M.A.J.,Van der Paauw, C.G., Tuinstra, L.G.M.T., Sauer, P.J.J., 1996.Effects of polychlorinated biphenyl dioxin exposure and feedingtype on infants mental and psychomotor development. Pediatrics97, 700–706.


Krieger, N., Wolff, M.S., Hiatt, R.A., Rivera, M., Vogelman, J.,Orentreich, N., 1994. Breast cancer and serum organochlorines –a prospective study among white, black and Asian women. J.Natl. Cancer Inst. 86, 589–599.

Lahvis, G.P., Wells, R.S., Casper, D., Via, C.S., 1993. In vitrolymphocyte response of bottlenose dolphins (Tursiops truncatus):mitogen induced proliferation. Marine Environ. Res. 35, 115–119.

Loganathan, B.G., Kannan, K., 1994. Global organochlorine con-tamination trends: an overview. Ambio 23, 187–191.

Lopez-Martın, J., Ruiz-Olmo, J., Minano, S.P., 1994. Organochlorineresidue levels in the European mink (Mustela lutreola) in North-ern Spain. Ambio 23, 294–295.

Lukoyanov, N.F., Zhukov, G.P., Meshkova V.G., Golovko, R.G.,1992. Experimental investigation of polydispersed solid aerosoldeposition on the snow surface. IEM Proc., 55(155), pp. 36–50 (inRussian).

Lund, J., Brandt, I., Poellinger, L., Bergman, A., Klasson-Wehler, E.,Gustafsson, J.-A., 1985. Target cells for the polychlorinatedbiphenyl metabolite 4,4%-bis(methylsulfonyl)-2,2%,5,5%-tetrachloro-biphenyl. Characterization of high affinity binding in rat andmouse lung cytosol. Mol. Pharmacol. 27, 314–323.

Lund, B.O., Bergman, A., Brandt, I., 1988. Metabolic activation andtoxicity of a DDT-metabolite, 3-methylsulphonyl-DDE, in theadrenal zona fasciculata in mice. Chem. Biol. Interactions 65,25–40.

Luster, M.I., Munson, A.E., Thomas, P.T., Holsapple, M.P., Fenters,J.D., White, K.L., Lauer, L.D., Germolec, D.R., Rosenthal, G.J.,Dean, J.H., 1988. Methods evaluation: development of a testingbattery to assess chemical-induced immunotoxicity – NationalToxicology Program’s guidelines for immunotoxicity evaluationin mice. Fund. Appl. Toxicol. 10, 2–19.

Luster, M.I., Portier, C., Pait, D.G., Rosenthal, G.J., Germolec,D.R., 1995. Immunotoxicology and risk assessment. In: Burleson,G.R., Dean, J.H., Munson, A.E. (Eds.), Methods in Immunotox-icology, vol. 1. Wiley-Liss, NY, pp. 51–68.

Mac, M.J., Edsall, C.C., 1991. Environmental contaminants and thereproductive success of lake trout in the Great Lakes: an epidemi-ological approach. J. Toxicol. Environ. Health 33, 375–394.

Mackay, D, Wania, F., 1995. Transport of contaminants to theArctic: partitioning, processes and models. Sci. Tot. Environ.160/161, 25–38.

Martineau, D., Beland, P., Desjardins, C., Lagace, A., 1987. Levels oforganochlorine chemicals in tissues of beluga whales (Delphi-napterus leucas) from the St. Lawrence estuary, Quebec, Canada.Arch. Environ. Contam. Toxicol. 16, 137–147.

Martineau, D., Lagace, A., Beland, P., Higgins, E., Armstrong, S.,Shugart, L.R., 1988. Pathology of stranded beluga whales (Del-phinapterus leucas) from the St. Lawrence estuary, Quebec,Canada. J. Comp. Path. 98, 287–311.

Mason, C.F., 1989. Water pollution and otter distribution: a review.Lutra 2, 97–131.

McConnachie, P.R., Zahalsky, A.C., 1991. Immunological conse-quences of exposure to pentachlorophenol. Arch. Environ. Health46, 249–253.

Mes, J., 1987. PCBs in human populations. In: Waid, J.S. (Ed.),PCBs and the Environment, vol. III. CRC Press, Boca Raton, FL,pp. 39–63.

Mocarelli, P., Brambilla, P., Gerthoux, P.M., Patterson Jr, D.G.,Needham, L.L., 1996. Change in sex ratio with exposure todioxin. The Lancet 348, 409.

Moore, M., Mustain, M., Daniel, K., Chen, I., Safe, S., Zacharewski,T., Gillesby, B., Joyeux, A., Balaguer, P., 1997. Antiestrogenicactivity of hydroxylated polychlorinated biphenyl congeners iden-tified in human serum. Toxicol. Appl. Pharmacol. 142, 160–168.

Morse, D.C., Klasson-Wehler, E., Wesseling, W., Koeman, J.H.,Brouwer, A., 1996. Alterations in rat-brain thyroid-hormonestatus following prenatal and postnatal exposure to polychlori-

nated-biphenyls (aroclor-1254). Toxicol. Appl. Pharmacol. 136,269–279.

Muir, D.C.G., Norstrom, R.J., Simon, M., 1988. Organochlorinecontaminants in Arctic marine food chains: accumulation ofspecific polychlorinated biphenyls and chlordane-related com-pounds. Environ. Sci. Technol. 22, 1070–1078.

Muir, D.C.G., Ford, C.A., Grift, N.P., Metner, D.A., Lockhart,W.L., 1990. Geographic variation of chlorinated hydrocarbons inburbot (Lota lota) from remote lakes and rivers in Canada, Arch.Environ. Contam. Toxicol. 19, 530–542.

Muir, D.C.G., Wagemann, R., Hargrave, B.T., Thomas, D.J.,Peakall, D.B., Norstrom, R.J., 1992. Arctic marine ecosystemcontamination. Sci. Tot. Env. 122, 75–134.

Muir, D.C.G., Grift, N.P., Lockhart, W.L., Wilkinson, P., Billeck,B.N., Brunskill, G.J., 1995. Spatial trends and historical profilesof organochlorine pesticides in Arctic lake sediments. Sci. Tot.Environ. 160/161, 447–457.

Mulvad, G., Pederson, H.S., Hansen, J.C., Dewailly, E., Jul, E.,Pedersen, M.B., Bjerregaard, P., Malcom, G.T., Deguchi, Y.,Middaugh, J.P., 1996. Exposure of Greenlandic Inuit toorganochlorines and heavy metals through the marine food-chain:an international study. Sci. Tot. Env. 186, 137–139.

Murray, F.J., Smith, F.A., Nitschke, K.D., Humiston, C.G., Kociba,R.J., Schwetz, B.A., 1979. Three-generation reproduction study ofrats given 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in the diet.Toxicol. Appl. Pharmacol. 50, 241–252.

Nagayama, J., Nagayama, M., Masuda, Y., 1992. Genotoxicity ofhighly toxic organochlorine congeners in cultured humanlymphocytes, Dioxin ’93: 13th International Symposium on Chlo-rinated Dioxins and Related Compounds, in: Fielder, H., Frank,H., Hutzinger, O., Parzefall, W., Riss, A., Safe, S. (Eds.). (FederalEnvironmental Agency, Vienna, September 1993) OrganohalogenCompounds, Vol. 13, pp. 165–167.

Norstrom, R.J., Simon, M., Muir, D.C.G., Schweinsberg, R.E., 1988.Organochlorine contaminants in Arctic marine food chains: iden-tification, geographical distribution and temporal trends in polarbears. Environ. Sci. Technol. 22, 1063–1071.

Olsson, M., Sandegren, F., 1991. Otter survival and toxic chemicalimplications for otter conservation programmes. In: Reuther, C.,Rochert, C. (Eds.), Proc. of the V Intern. Otter ColloquiumHabitat 6. Hankensbuttel, Germany, pp. 191–200.

Olsson, M., Bignert, A., Odsj, T., Persson, W., Litzen, K., Eriksson,U., Haggberg, L., Alsberg, T., 1997. Temporal trends oforganochlorines in Northern Europe, 1967-1995 support longranged transport but not the ‘grass-hopper effect’. In: R. Hites(Ed.), Dioxin ’97, Organohalogen Compounds, Vol 33, pp. 99–104.

Osterhaus, A.D.M.E., Vedder, E.J., 1988. Identification of viruscausing recent seal deaths. Nature 335, 20.

Osterhaus, A.D.M.E., De Swart, R.L., Ross, P.S., Timmerman,H.H., Van Loveren, H., Reijnders, P.J.H., Vos, J.G., 1995. Im-pairment of immune function in harbour seals (Phoca 6itulina)feeding on fish contaminated through the food chain, Abstractpresented at the Conference. Chemically-induced Alterations inthe Developing Immune System: The Wildlife/Human Connec-tion, Racine, Wisconsin, February 10-12, 1995.

Ottar, B., 1981. The transfer of airborne pollutants to the Arcticregions. Atmos. Environ. 15, 1439–1445.

Ouellet, M., Bonin, J., Rodrigue, J., DesGranges, J., Lair, L., 1997.Hind limb deformities (ectromelia, ectrodactyly) in free livinganurans from agricultural habitats. J. Wildl. Dis. 33, 95–104.

Perera, F., 1981. Carcinogenicity of airborne fine particulate ben-zo(a)pyrene: An appraisal of the evidence and the need forcontrol. Environ. Health Perspect. 42, 163–185.


Peterson, B.J., Fry, A.J., 1987. Stable isotopes in ecosystem studies.Annu. Rev. Ecol. Syst. 18, 293–320.

Peterson, R.E., Moore, R.W., Mably, T.A., Bjerke, D.L., Goy, R.W.,1992. Male reproductive system ontogeny: effects of perinatalexposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin in Chemically-In-duced Alterations. In: Colborn, T., Clement, C. (Eds.), Sexualand Functional Development: The Wildlife/Human Connection.Princeton Scientific Publishing, Princeton, NJ, pp. 175–193.

Quaß, U., Fermann, M., 1997. Identification of relevant industrialsources of dioxins and furans in Europe, (The European DioxinInventory). Final Report, Materialen No. 43, published by NorthRhine-Westphalia State Environment Agency (LUA NRW)),ISSN 0947-5206.

Rahn, K.A., Heidam, N.Z., 1981. Progress in Arctic air chemistry1977-1980: a comparison of the first and second symposium.Atmos. Environ. 15, 1345–1348.

Rappe, C., 1974. Chemical behavior of pesticides. In: Bylund, E.,Linderholm, H., Rune, O. (Eds.), Ecological Problems of theCircumpolar Area. Norrbottens Museum, Lulea, Sweden, pp.29–32.

Reijnders, P.J.H., 1986. Reproductive failure in common seals feedingon fish from polluted coastal waters. Nature 324, 456–457.

Renn, O., 1995. Style of using scientific expertise: a comparativeframework. Sci. Public Policy 22, 147–156.

Renn, O., Webler, T., Wiedemann, P., 1995. Fairness and Compe-tence in Citizen Participation. Evaluating Models for Environ-mental Discourse. Kluwer Academic Publishers, Dordrecht.

Repetto, R., Baliga, S.S., 1996. Pesticides and the immune system:The public health risks. World Resources Institute, Washington,DC, USA.

Risebrough, R.W., Carmignani, G.M., 1972. Chlorinated hydrocar-bons in Antarctic birds. In: Parker, B.C. (Ed.), Proceedings of thecolloquium, Conservation problems in Antarctica. Allen Press,Lawrence, Kansas, pp. 63–78.

Ritter, L., Soloman, K.R., Forget, J., Stemeroff, M., O’Leary, C.,1995. Persistent organic pollutants, an assessment report on DDT,Aldrin, Dieldrin, Endrin, Chlordane, Heptachlor, Hexachloroben-zene, Mirex, Toxaphene, Polychlorinated biphenyls, Dioxins andFurans. Report No. PCS/95.38 prepared for the InternationalProgramme on Chemical Safety (IPCS) within the framework ofthe Inter-Organization Programme for the Sound Management ofChemicals (IOMC). (http://irptc.unep.ch/pops/indxhtms/asses0.html)

Rogan, W., Gladen, B., McKinney, J., Carreras, N., Hardy, P.,Thullen, J., Tingelstad, J., Tully, M., 1986. Neonatal effects oftransplacental exposure to PCBs and DDE. J. Pediatr. 109,335–341.

Rola, A.C., Pingali, P.L., 1993. Pesticides, rice production and farm-ers. International Rice Research Institute and World ResourcesInstitute, Washington, DC.

Ross, P.S., De Swart, R.L., Reijnders, P.J.H., Van Loveren, H., Vos,J.G., Osterhaus, A.D.M.E, 1995. Contaminant-related suppres-sion of delayed-type hypersensitivity and antibody responses inharbor seals fed herring from the Baltic Sea. Environ. HealthPerspec. 103, 162–167.

Safe, S.H., 1994. Polychlorinated-biphenyls (PCBs)-environmental-impact, biochemical and toxic responses, and implications for riskassessment. Crit. Rev. Toxicol. 24, 87–149.

Safe, S.H., 1997. Is there an association between exposure to environ-mental estrogens and breast cancer? Environ. Health Perspec. 105,675–678.

Segstro, M., Muir, D., Hobson, K., Stewart, R., Olpinski, S., 1993.Are unexpectedly high levels of PCBs and other organochlorinesin walrus due to predation on seals? Presented at the InternationalConference on Marine Mammal Biology, Galveston, Texas,November 1993.

Seys, A., 1997. Dioxins-related compounds in the chemical industry-Presentation to the UNEP sub-regional meeting on identificationand assessment of releases of persistent organic pollutants, St.Petersburg, Russian Federation, 1-4 July 1997. UNEP, Geneva.(http://irptc.unep.ch/pops/stpeter/stpete13.html)

Sicre, M.A., Marty, J.C., Saliot, A., Aparicio, X., Grimalt, J., Al-baiges, J., 1987. Aliphatic and aromatic hydrocarbons in Mediter-ranean aerosol. Int. Environ. Anal. Chem. 29, 73–94.

Simonich, S.L., Hites, R.A., 1995. Global distribution of persistentorganochlorine compounds. Science 269, 1851–1854.

Sjoberg, L., 1995. Explaining risk perception: An empirical andquantitative evaluation of cultural theory (RHIZIKON: RiskResearch Reports No. 22). Centre for Risk Research, StockholmSchool of Economics, Stockholm.

Sladen, W.L.J., Menzie, C.M., Reichel, W.L., 1966. DDT residues inAdelie penguins and a crabeater seal from Antarctica. Nature 210,637–670.

Svensson, B.G., Hallberg, T., Nilsson, A., Akesson, B., Schutz, A.,Hagmar, L., 1993. Immunological competence and liver functionin subjects consuming fish with organochlorine contaminants,Dioxin ’93: 13th International Symposium on Chlorinated Diox-ins and related compounds, Federal Environmental Agency, Vi-enna, September 1993, Organohalogen Compounds, Vol. 13,pp.175–178.

Swain, W., Colborn, T., Bason, C., Howarth, R., Lamey, L., Palmer,B., Swackhamer, D., 1992. Chapter 22: Exposure and effects ofairborne contamination. The Great Waters Program Report.United States Environmental Protection Agency, USA.

Tanabe, S., Hidaka, H., Kawano, M., Tatsukawa, R., 1982. Globaldistribution and atmospheric transport of chlorinated hydrocar-bons: HCH isomers and DDT compounds in the Western Pacific,Eastern Indian and Antarctic Ocean. J. Oceanog. Soc. Jpn. 5,97–109.

Tattersfield, L., Matthiessen, P., Campbell, P., Grandy, N., Lange,R., 1997. SETAC-Europe/OECD/EC Expert Workshop on En-docrine Modulators and Wildlife: Assessment and Testing(EMWAT), held at Veldhoven, The Netherlands, 10–13 April,1997. SETAC-Europe, Brussels.

Thomann, R.V., Connolly, J.P., Parkerton, T.F., 1992. An equi-librium model of organic chemical accumulation in aquatic foodwebs with sediment interaction. Environ. Toxicol. Chem. 11,615–629.

Thomas, D.J., Tracy, B., Marshall, H., Norstrom, R.J., 1992. Arcticterrestrial ecosystem contamination. Sci. Tot. Environ. 122, 135–164.

Thomas, K.B., Colborn, T., 1992. Organochlorine endocrine dis-rupters in human tissue. In: Colborn, T., Clement, C. (Eds.),Chemically-Induced Alterations in Sexual and Functional Devel-opment: The Wildlife/Human Connection. Princeton ScientificPublishing, Princeton, NJ, pp. 365–394.

Toppari, J., Larsen, J.C., Christiansen, P., Giwercman, A., Grand-jean, P., Guillette, L.J., Jegou, B., Jensen, T.K., Jouannet, P.,Kieding, N., Leffers, H., McLachlan, J.A., Meyer, O., Muller, J.,Rajpert-De Meyts, E., Scheike, T., Sharpe, R., Sumpter, J.,Skakkebaek, N.E., 1996. Male reproductive health and environ-mental xenoestrogens. Environ. Health Perspect. 104, 741–803.

Tryphonas, H., Luster, M.I., Schiffman, G., Dawson, L.L., Hodgen,M., Germolec, D., Hayward, S., Bryce, F., Loo, J.C.K., Mandy,F., Arnold, D.L., 1991. Effect of chronic exposure of PCB(Aroclor 1254) on specific and non-specific immune parameters inthe rhesus (Macaca mulatta) monkey. Fund. Appl. Toxicol. 16,773–786.

UN-ECE, 1994. State of knowledge report of the UN ECE task forceon persistent organic pollutants (Draft, April 1994). For theUN-ECE Convention on Long-range Transboundary AirPollution.


UN-ECE, 1998a. Draft Protocol to the Convention on Long-rangeAir Pollution on Persistent Organic Pollutants, (EB.AIR/1998/2),Convention on Long-range Transboundary Air Pollution, UnitedNations Economic and Social Council, Economic Commission forEurope.

UN-ECE, 1998b. Draft Executive Body Decision 1998/2 on informa-tion to be submitted and the procedure for adding substances toAnnexes I, II or II to the POPs Protocol, Convention on Long-range Transboundary Air Pollution, United Nations Economicand Social Council, Economic Commission for Europe.

UNEP, 1996. UNEP survey on sources of POPs, A report preparedfor an IFCS expert meeting on persistent organic pollutants,Manila, the Philippines, 17–19 June 1996, UNEP, Geneva. (http://irptc.unep.ch/pops/indxhtms/manexp3.html)

Van Birgelen, A.P.J.M., Fase, K.M., van der Kolk, J, Poiger, H.,Brouwer, A., Seinen, W., van den Berg, M., 1996. Synergisticeffect of 2,2%,4,4%,5,5%-hexachlorobiphenyl and 2,3,7,8-tetra-chlorodibenzo-p-dioxin on hepatic porphyrin levels in the rat.Environ. Health Perspect. 104, 550–557.

Van Jaarsveld, J.A., van Pul, W.A.J., de Leeuw, F.A.A.M., 1997.Modelling transport and deposition of persistent organic pollu-tants in the European region. Atmos. Environ. 31, 1011–1024.

Van Leeuwen, C.J., Bro-Rasmussen, F., Feijtel, T.C.J., Arndt, R.,Bussian, B.M., Calamari, D., Glynn, P., Grandy, N.J., Hansen,B., Van Hemmen, J.J., Hurst, P., King, N., Koch, R., Muller, M.,Solbe, J.F., Speijers, G.A.B., Vermeire, T., 1996. Risk assessmentand management of new and existing chemicals. Environ. Toxicol.Pharmacol. 2, 243–299.

Van Velsen, F.L., Danse, L.H.J.C., Van Leeuwen, F.X.R., Dormans,J.A.M.A., Van Logten, M.J., 1986. The subchronic oral toxicityof the beta-isomer of hexachlorocyclohexane in rats. Fund. Appl.Toxicol. 6, 697–712.

Vozhzennikov, O.J., Bulgakov, A.A., Popov, V.E., Lukoyanov, N.F.,Naidenov, A.V., Burkov, A.J., Kutnyakov, J.V., Hzirnov, V.G.,1997. Review of migration and transformation parameters forselected POPs, EMEP/MSC-E Report 3/97. Meteorological Syn-thesizing Centre East, Russia.

Wania, F., 1997. Obstacles to deriving critical loads for persistentorganic pollutants. Poster presentation at the Workshop on Criti-cal Limits and Effects Based Approaches for Heavy Metals andPersistent Organic Pollutants. Bad Harzburg, November 3-7, Ger-many, p. 1997.

Wania, F, Mackay, D., 1993. Global fractionation and cold conden-sation of low volatility organochlorine compounds in Polar Re-gions. Ambio 22, 10–18.

Wania, F, Mackay, D., 1996. Tracking the distribution of persistentorganic pollutants. Environ. Sci. Technol. 30, 390–396.

Weber, K., Goerke, H., 1996. Organochlorine compounds in fish offthe Antarctic Peninsula. Chemosphere 33, 377–392.

Weisglas-Kuperus, N., Sas, T.C.J., Koopman-Esseboom, C., van derZwan, C.W., de Ridder, M.A.J., Keishuizen, A., Hooijkaas, H.,Sauer, P.J.J., 1995. Immunological effect of background prenataland postnatal exposure to dioxins and polychlorinated biphenylsin Dutch infants. Pediatr. Res. 38, 404–410.

Weschler, C.J., 1981. Identification of selected organics in the arcticaerosol. Atmos. Environ. 15, 1365–1369.

WHO, 1996, Persistent organic pollutants, WHO, Geneva.Wolfe, D., Marquardt, H., 1993. The tumor promoting potential of

plychlorinated dioxins and biphenyls: mechanisms of action indifferent in vitro assays, in: Fielder, H., Frank, H., Hutzinger, O.,Parzefall, W., Riss, A., Safe, S. (Eds.), Dioxin ’93: 13th Interna-tional Symposium on Chlorinated Dioxins and Related Com-pounds, (Federal Environmental Agency, Vienna, September1993) Organohalogen Compounds, Vol. 13, pp. 163–166.

Wolff, M.S., Toniolo, P.G., Lee, E.W., Rivera, M., Dubin, N., 1993.Blood levels of organochlorine residues and risk of breast cancer.J. Nat. Cancer Inst. 85, 648–652.

Wren, C.D., 1991. Cause-effect linkages between chemicals and popu-lations of mink (Mustela 6ision) and otter (Lutra canadensis) inthe Great Lakes basin. J. Toxicol. Environ. Health 33, 549–586.

Wynne, B., 1996. Misunderstood Misunderstandings: social identitiesand public uptake of science. In: Irwin, A., Wynne, B. (Eds.),Misunderstanding Science? The public reconstruction of scienceand technology. Cambridge University Press, Cambridge.

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