Download - Contaminants and their effects on estuarine and coastal organisms in the United Kingdom in the late twentieth century

Contaminants and their effects on estuarine and coastal organismsin the United Kingdom in the late twentieth century

Peter Matthiessen*, Robin J. Law

Centre for Environment, Fisheries and Aquaculture Science, Remembrance Avenue, Burnham-on-Crouch, Essex CM0 8HA, UK

Received 31 May 2001; accepted 28 February 2002

‘‘Capsule’’: The entire field of contaminant effects in UK estuarine and coastal waters is reviewed, focusing on recentyears and recommendations for monitoring/research.

Abstract

The biological effects of contaminants in British estuaries and coastal waters have been studied for over 100 years. Until the1970s, the major pollution impact on estuarine organisms was probably caused by poorly treated sewage which led to severe oxygen

deficits and consequent asphyxiation of many water-breathers. However, since the introduction of improved sewage treatment inthe last 30 years, a number of continuing impacts have come to light which represent true toxic effects of micro-contaminants.Sublethal changes observed in various bioassays and biomarkers are widespread in both fish and invertebrates, and in the most

urbanised and industrialised estuaries, these effects are probably having impacts at the population and community levels. Forexample, there is good evidence to show that tributyltin from some antifouling paints has not only affected the sexuality andreproductive success of individual estuarine and coastal molluscs, but has also damaged some benthic communities of which mol-

luscs are but a part. Although there are data to show that some contaminant concentrations are now declining, we do not yet haveenough data in most cases to decide whether organisms have also begun to recover. This paper reviews the entire field of con-taminant effects in UK estuarine and marine waters, focusing especially on the years 1985–2000, and makes some recommendationsfor future research and monitoring programmes. Published by Elsevier Science Ltd.

Keywords: Contaminants; Biological effects; Marine environment; Monitoring; United Kingdom

1. Introduction

The effects of pollution on estuarine organisms in theUnited Kingdom have been of concern in some areas forwell over 100 years. One of the earliest investigationswhich specifically addressed this issue was described byAlbert Gunther of the BritishMuseum (Natural History)in London in a report to the Metropolitan Sewage Dis-charge Commission (Gunther, 1883, cited in Wheeler,1979). He was responding to concerns that the construc-tion of the Northern and Southern Outfall sewers, whichdischarged untreated domestic sewage from much of theLondon area into the Thames estuary at Barking Reach,may have been damaging important fisheries. There werealso concerns that the ammoniacal effluent from Beckton

gas-works (producing coal gas), the largest such worksin Europe at that time, was poisoning organisms in theadjacent Gallions Reach. These concerns were realenough: few if any resident fish species were present inthe relevant part of the Thames estuary, and severalmigratory species were unable to move upstream.Gunther tackled these questions by using inter alia

what would now be called in situ bioassays. He placedfish (eels and flounder) and crustaceans (shrimps) inperforated fish boxes in the estuary, as well as in bucketscontaining various dilutions of estuarine water andeffluents, and observed the ensuing mortality. The majorconclusion was that discharges of sewage effluent wereleading to reduced dissolved oxygen concentrationswhich then caused asphyxia of water-breathing organ-isms, especially during summer. However, some con-stituents of gas-works effluent (ammonia, cyanide,sulphide, phenols, etc.) were also identified as having aspecific toxic effect. Many kilometres of the middleThames estuary were found to be unfit for fish and/or

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Environmental Pollution 120 (2002) 739–757

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* Corresponding author. Present address: Centre for Ecology and

Hydrology, Far Sawrey, Ambleside, Cumbria LA22 0LP, UK. Fax:

+44-15394-46914.

E-mail address: [email protected] (P. Matthiessen).

shrimps, although the waters upstream of LondonBridge were deemed of satisfactory quality.These observations led directly to the introduction of

primary sewage treatment to remove suspended solidsfrom about 1884–1888 onwards. This in turn led to anincrease of dissolved oxygen near the major outfalls toabout 25% saturation in 1905, and thence to the returnof some fish species (e.g. flounder, smelt and whitebait)to the Thames estuary (Water Pollution ResearchLaboratory, 1964; Port of London Authority, 1967).However, the increasing population of London, produ-cing ever more sewage, was the cause of a reversal inthis recovery process, such that by 1920, oxygensaturation near the major discharges had declined tojust 5%. Large stretches of the estuary remained anae-robic (and therefore fishless) until after the SecondWorld War, when new treatment plant began to banishunacceptably low dissolved oxygen levels from the mid-1960s onwards. Wheeler (1979) documented the sub-sequent recovery of the Thames fish fauna, and showedthat whereas in 1957 there were no established fishpopulations for about 69 km below Kew Bridge inLondon, by 1967–1973 a total of 72 fish species could becollected from the estuary (18 freshwater species, 43marine, 11 euryhaline). An example of the rapidity ofthis recovery is given by the flounder (Platichthys flesus),of which only two specimens were caught on powerstation intake screens in 1967, while 139 were capturedby 1972 from all stretches of the tidal river. Fish com-munities in the Thames estuary have remained broadlystable and even increased in species since the 1970s, with19 freshwater and 92 marine species being recorded bythe 1990s (Andrews, 1984; Araujo et al. 2000; EA,1997), and a total of 118 species reported to be presentin the river as a whole in 2001 (Anon, 2001).The situation in other British estuaries was not recor-

ded in this degree of detail (e.g. Meek, 1923; Jørgensen,1929; Alexander et al., 1935; Gascoigne and Wildish,1971; Porter, 1973; Mackay et al., 1978; Rees andEleftheriou, 1989), but the Thames was certainly notunrepresentative of other estuaries with poor waterexchange that flowed through centres of human popu-lation and industry (e.g. Clyde, Forth, Mersey, Tees,Tyne, Humber, Medway, Swale). In essence, inputs oforganic matter had exerted such a huge biological oxy-gen demand that fish and other organisms were severelydepleted or absent in such estuaries for much of theirlength, largely due to the anaerobic conditions. Some ofthe most detailed data outside the Thames wereobtained on fish populations in the Forth (Webb andMetcalfe, 1987; Elliott et al., 1988), where annual sum-mer oxygen deficits in the fresh–brackish transition zonepersisted into the 1980s and caused salmon kills andimpaired migration in salmonids and flounder.When considering the recent history of contaminants

and their effects in UK estuaries, it is clearly important

to bear in mind the facts of earlier gross organic pollu-tion described above. In several cases, aquatic specieshave only been re-colonising major estuaries for 30years. This in turn means that populations may nothave had sufficient time to evolve resistance to somepollutants, and sublethal effects of trace contaminantson such vital processes as growth and reproduction maynow be expressed which formerly were irrelevant due tothe more or less complete absence of viable aquaticcommunities.It is also important to consider the legislative changes

which have brought about some reductions in dis-charges and improved environmental quality. In fact,legal controls over direct sewage and industrial dis-charges to UK coastal waters and estuaries were onlyestablished relatively recently, in comparison with muchearlier regulation of discharges to rivers. The first con-trols were incorporated in the Clean Rivers (Estuariesand Tidal Waters) Act 1960, imposing a requirement forconsents on new discharges, but existing land-baseddischarges to tidal waters (out to the 3 mile, 5 km, limit)were only brought under regulation in the Control ofPollution Act 1974. This made it an offence to ‘cause orknowingly permit any poisonous, noxious or pollutingmatter to enter controlled waters’. A plethora of Eur-opean Community Directives has led to further controlson discharges to estuarine and coastal waters (e.g. theDangerous Substances Directive 74/464/EEC anddaughter directives, which inter alia set environmentalquality standards for a number of substances; theShellfish Waters Directive 79/923/EEC which specified arange of further standards for waters supporting shell-fish; and the Urban Waste Water Treatment Directive91/271/EEC which specified the degree of treatment tobe given various sewage discharges). More recently, theEnvironment Protection Act 1990 and the EC Directiveon Integrated Pollution, Prevention and Control (96/61/EEC) introduced the concept of Integrated PollutionControl in order to ensure that the effect of any releaseto the environment is minimised. Finally, the ScottishEnvironmental Protection Agency and the England andWales Environment Agency were brought into being bythe Environment Act 1995, and are responsible forapplying the various controls on discharges to estuariesand coastal waters.In offshore waters, disposal of wastes (now limited to

dredged materials and some fish waste) is under the con-trol of the UK Department for Environment, Food andRural Affairs (DEFRA—formerly the Ministry of Agri-culture, Fisheries and Food, MAFF) through the Foodand Environment Protection Act 1985, and is subject to arigorous assessment of environmental impact. The onlytypes of polluting discharges to have escaped regulationto date concern inputs of substances from offshore dril-ling rigs and platforms, but these are now covered by theOffshore Chemicals Regulations 2001 administered by

740 P. Matthiessen, R.J. Law /Environmental Pollution 120 (2002) 739–757

the UK Department of Trade and Industry under thePollution Prevention and Control Act 1999.This paper aims to review the current status of the

presence and effects of trace contaminants in UKestuarine and coastal waters, as exemplified by datasetsgathered by CEFAS and others over the last 10 to 15years. Much of the information derives from theNational Monitoring Programme (NMP)–now theNational Marine Monitoring Programme (NMMP).The first report of the NMP (NMP, 1998) makes gen-erally optimistic reading. Fisheries and wildlife do notappear to be in serious decline due to contaminanteffects, and many contaminant concentrations areapparently decreasing. Furthermore, estuary quality asdefined by the rather crude National Water Councilestuary classification scheme (based on dissolved oxygenconcentrations, aesthetic quality and biological qual-ity—EA, 1999) has been improving since the 1980s, andby 1995 there were only nine estuaries which containedone or more stations classified as ‘bad’ or ‘poor’. Thiscorresponded to an estuarine length of 223 km, or 8%of the total.There is little doubt that the major human impact on

marine ecosystems in the North Sea is now caused bythe effects of fishing, partly due to the removal of hugetonnages of fish (30–40% of the biomass of exploitedspecies in the North Sea is removed annually), andpartly because of unintentional damage. For example, arecent paper has shown that diversity of epibenthic spe-cies across large areas of the North Sea is negativelycorrelated with beam trawling effort (Zuhlke et al.,2001) and it has been argued that the impacts nowcaused by contaminants are small in comparison. Thisview was borne out both by a review of the coastalenvironment by the England and Wales EnvironmentAgency (EA, 1999), and by a comprehensive review ofpollution impacts on North Sea fish stocks (Parrett,1998) published by WWF, UK.However, it will become apparent that trace or micro

contaminants are still causing demonstrable pollutingeffects in areas where certain marine discharges arepoorly diluted, or where there is a legacy of persistentpollutants in fine-grained sedimentary sinks. The NMPreport confined itself to mentioning only a few biologi-cal measures of pollution, but stated that contaminantsare significantly stressing several species, especially inestuaries, with acutely toxic waters being present insome. The big question is whether any of these effectsare causing damage to populations and communities,because effects confined to individuals alone are notusually considered to be a threat to ecosystem structureor function (McIntyre and Pearce, 1980; Gray, 1989). Itwas also mentioned by the NMP that environmentalquality standards (EQS) in UK marine waters are gen-erally being met, but this of course does not necessarilyimply that problems do not exist because it ignores the

possible effects of mixtures, and the fact that no EQSvalues have been developed for the overwhelmingmajority of anthropogenic marine contaminants.

2. Trends in, and possible significance of, contaminant

levels in the UK coastal and estuarine environment

There is no doubt that the major improvements inestuarine and river quality seen over the past centuryhave occurred as a result of the reduction in dischargesof raw sewage to watercourses. As has been mentionedabove, the earliest amelioration efforts began in theRiver Thames. The initial capital works, involving theconstruction of interceptor sewers and the pumping ofthe collected sewage streams to treatment works atBeckton and Crossness, downstream of London, werecompleted in 1865 (Casapieri, 1984). Similar schemeshave followed in other major rivers, including the Mer-sey (Olsen et al., 1999) and Tyne (Norgrove, 1977), andin some coastal areas where there have been difficultiesin meeting bathing water quality standards (e.g. thecoast of NW England, Atkinson, 1997). More generallythe aim has been to satisfy environmental qualityobjectives initially set to ensure the passage of migratoryfish, requiring for the highest standard of freshwaterriver (1A) dissolved oxygen saturation above 80%, bio-logical oxygen demand equal to or less than 3 mg/l,ammonia concentrations below 0.4 mg/l, and that‘‘visible evidence of pollution should be absent’’ (Casa-pieri, 1984). Subsequently this was superseded by ageneral quality assessment protocol (NRA, 1994).In at least one instance improvements in water quality

have led to unforeseen problems. When fish began toreturn to the inner Mersey estuary the elevated con-centrations of some contaminants (e.g. DDT and itsmetabolites, PCBs and mercury) observed in their tis-sues raised concern of possible risks to the health ofhuman consumers, and advisory notices were issued toanglers (Collings et al., 1996; Leah et al., 1997 a, b). ForPCBs and other previously widely used industrial andagricultural organochlorines (such as DDT, g-HCH,HCB and dieldrin) contamination can be observed in allmajor estuaries, with variations in concentrationbetween localities. Study of dated saltmarsh sedimentcores has demonstrated the rise and fall of inputs of anumber of organochlorine compounds resulting fromthe cycle of production, use and legislative controls(Fox et al., 2001). Fig. 1 illustrates this for PCBs (as thesum of the seven congeners specified by the Interna-tional Council for the Exploration of the Sea) in sedi-ments from Banks Marsh in NW England. PCBcontamination became apparent in the cores from thelate 1940s, rose to a peak about 1970 and has declinedsubsequently following the cessation of production andrestrictions upon use and discharge.

P. Matthiessen, R.J. Law /Environmental Pollution 120 (2002) 739–757 741

Individual estuaries often exhibit contaminationresulting from specific uses and processes undertakenwithin their catchments. In the Mersey estuary, forinstance, there were considerable inputs of mercury dueto the operation of chlor-alkali plants, although theseinputs have been drastically reduced over time. In thecase of one such factory situated at Runcorn, thisreduction was from 60 tpa in the 1970s to <1 tpa by1993 (NRA, 1995a). As a result of these historical dis-charges, however, there remains an accumulated reser-voir of mercury in the sediments of the estuary andinner Liverpool Bay, as evidenced by the analysis ofdated saltmarsh sediment cores. Concentrations in fishfrom the area also exhibit elevated concentrations ofmercury relative to other areas of the UK, althoughthey have long been in decline as a result of the reduc-tions in inputs (DETR, 2000; see Fig. 2). Spatial pat-terns of mercury contamination in fish can also beobserved throughout the NE Irish Sea, with decliningconcentrations related to their distance from the Merseyestuary (Leah et al., 1991). Biota of the Mersey estuaryare also contaminated by a number of isomers ofmethylchlorocyclohexanes arising from a specific indus-trial process in the area (McNeish et al., 1999). In gen-eral, however, organochlorine contamination (includingmethychlorocyclohexanes) in Mersey mussels (Mytilusedulis) was declining between 1994 and 1998 (Connor etal., 2001).

Similarly, sediments and biota in the Tees estuaryhave been contaminated with brominated diphenylethers manufactured and used within the catchment(Allchin et al., 1999); concentrations of alkylphenolethoxylates and their nonylphenol degradation producthave been elevated in the Rivers Aire, Mersey and Tees(Blackburn and Waldock, 1995; Blackburn et al., 1999);concentrations of a range of volatile organic com-pounds (e.g. carbon tetrachloride, chloroform, tri-chloroethene and tetrachloroethene) were elevated inthe Mersey estuary as a result of inputs from the Man-chester Ship Canal and the River Weaver (Rogers et al.,1992); and the use of HCH in the textile industry in theWest Riding of Yorkshire in the early 1980s led directlyto contamination of the River Aire downstream of theseactivities (NRA, 1993).In England and Wales the Environment Agency (EA)

has the primary responsibility for monitoring waterquality in rivers, estuaries, and coastal waters to 3 milesfrom the shoreline. Although some EA-funded workand research is aimed at establishing true environmentalconcentrations for contaminants, the main focus isdirected towards the statutory functions of the agency.In this context that is to undertake compliance mon-itoring, intended to establish whether the concentrationsof a range of contaminants fall below environmentalquality standard (EQS) values. These are set on thebasis of toxicity of each compound individually, and so

Fig. 1. PCB concentrations (as the sum of the ICES 7 congeners) in a dated saltmarsh sediment core from Banks Marsh, NW England (mg/kg dry

wt.; redrawn after Fox et al., 2001).


provide reassurance that a pesticide, for instance, is notpresent at harmful concentrations. Unfortunately thisneither demonstrates real concentration trends (as onceconcentrations fall below ca. 10% of the EQS value theyare generally below the limits of determination of themethods selected), nor takes any account of the toxicityof either mixtures or compounds for which an EQSvalue has not yet been established. As an example ofthis approach, Meharg et al. (1998) determined theconcentrations of a range of organochlorine con-taminants in the more southerly industrial rivers (Aire,Calder, Don and Trent) feeding the Humber estuary atleast once a week over an 18-month period. These datawere then assessed on the basis of EQS values for eachcompound, with more frequent exceedances of EQSvalues apparent in samples from the Aire and Calder. Ingeneral temporal trends have been addressed in terms ofloads discharged to the environment rather than inchanges in the environmental concentrations resultingfrom these discharges (e.g. NRA, 1995b). Flux estimatesmay exhibit strong variability both daily and seasonally(Zhou et al., 1998), and there are also difficulties inestimating loads for organic contaminants due to thenumber of instances in which concentrations are belowlimits of determination.Burt et al. (1992) compared metal contamination in a

number of UK estuaries by determining concentrationsin Fucus vesiculosus, Littorina littorea,Macoma balthica,Nereis diversicolor and Scrobicularia plana, so as toprepare a series of maps which would form the basis for

studies of future trends, as well as allow spatial com-parisons between estuaries. Long-term metal pollutionresulting from mining activities undertaken over a per-iod of two centuries was studied in the Fal estuary inSW England (Bryan et al., 1987). Analysis of copperand zinc in sediments indicated that although levels ofcopper from mining had declined, those of zinc had not,and that both metals were still exerting an influenceupon the fauna of the estuary. Temporal trends in metalpollution in the Severn estuary were studied by Allenand Rae (1986), who concluded that lead and zincentered the estuary largely from natural sources, butthat their fluxes were anthropogenically enhanced dur-ing the industrial and post-industrial periods (ca. 1840–1950). Furthermore, an excellent dataset on temporaltrends of heavy metals in mussels from the Forth estu-ary (Dobson, 2000) shows that mercury, cadmium andchromium levels declined markedly between 1981 and1999. On the other hand, Matthiessen et al. (1999) con-cluded that much of the copper and zinc contaminationin south-eastern estuaries, which is causing regularexceedances of EQS values, originates con-temporaneously from antifouling paints and sacrificialanodes, respectively, on small vessels. Concentrations ofPAH, metals and organochlorine contaminants in asediment core from Tilbury on the River Thamesshowed a marked decrease at a depth of 5.8 m, resultingfrom reduced inputs due to improvements in sewagetreatment in the late 1950s and early 1960s (Taylorand Lester, 1995). Power et al. (1999) studied trends in

Fig. 2. Time series of mean concentrations of mercury in fish flesh from Liverpool and Morecambe Bays (mg/kg wet wt.; DETR, 2000).


agricultural pesticide (atrazine, lindane, simazine) con-centrations in the Thames estuary between 1988 and1997 in order to determine whether attempted reduc-tions in pesticide discharges had improved water qual-ity. They found that concentrations of all the studiedpesticides had declined over the period, despite theinfluence of drought-induced reductions in freshwaterflow from the river catchment. The authors noted,however, that ‘‘little has been done [in the UK] to docu-ment either the spatial or temporal character of pesticidedischarges to estuarine environments. The work that hasbeen completed focuses mainly on the spatial aspects ofpesticide distribution within estuarine environments.’’This aspect is one that is currently being addressed

within the UK National Marine Monitoring Pro-gramme which, now that its initial spatial phase hasbeen completed (NMP, 1998), has switched focus toaddress temporal trends in contamination at selectedsites around the UK. Within this programme the con-taminant studies are integrated with both biologicaleffects measurements and benthic community studies.This will lead, in time, to a much better understandingof the changes in concentrations of environmental con-taminants to which organisms are exposed, but, ofnecessity, the programme will need to be responsive toinformation proceeding from research studies such asthose currently investigating endocrine disruption (e.g.Allen et al., 1999a, b) and other pollution problems inUK estuaries (Thomas et al., 1999a, b), so as to main-tain a relevant suite of determinands for the future.

3. Biological effects of contaminants in the UK coastal

and estuarine environment

3.1. Bioassays of water

One of the earliest modern biological measures ofwater quality applied to marine waters around the UKwas the oyster (Crassostrea gigas) embryo bioassay(Thain, 1991; CEFAS, 1998; NMP, 1998). This bioassaymeasures both lethal and sublethal toxicity in develop-ing embryos exposed to a water sample over a 24-hexposure period. It has been used routinely on low-tidesub-surface samples since the early 1990s, particularly inthe period 1990–1995, and gives a picture of variablebut recurring toxicity in the waters of several Englishestuaries, including the Tees, Tyne, Wear and Mersey.Sporadic toxicity has also been observed with thisbioassay in the estuaries of the Tweed, Blyth, Humber,Great Ouse, Tamar and Dee, and in Poole Harbour.High toxicity has not been observed in coastal waters,although slightly toxic samples have occasionally beenobtained immediately offshore from the Mersey,Thames, Tamar and Tyne estuaries, and in the centralEnglish Channel. Reasonably typical data are shown in

Table 1 for the year 1994, but marked inter-year varia-bility is a feature of many estuarine stations, and isillustrated by data for Redcar Jetty on the Tees whichrevealed high toxicity in May–July in the years 1990,1991, 1993 and 1994, but none in 1992.Insufficient oyster embryo bioassay data have yet

been collected to give firm information on temporaltrends, but water quality in industrialised estuariesdetermined by this bioassay does not appear to haveundergone major improvements during the 1990s.Water quality in coastal and offshore waters, with rareexceptions, is consistently favourable in short-term testswith oyster embryos, but it is of interest to know whe-ther contaminant concentrations in these areas arechronically toxic or are approaching acutely toxic levels.As the oyster bioassay and related tests are not sensi-

tive to the lower levels of contaminants found in watersoffshore, and in view of the fact that chronic bioassays(e.g. Stebbing et al., 1991) are too time consuming forgeneral survey work, some surveys have focused onhexane-extracted concentrates of seawater bioassayedusing the harpacticoid copepod Tisbe battagliai exposedfor 48 h (CEFAS, 1998; Kirby et al., 1998; Thain andKirby, 1996). This showed that non-polar organic con-taminant concentrations in several industrialised estu-aries and a harbour (estuaries of the Tyne, Wear, Teesand Mersey, and Poole Harbour) in July/October 1992and June/October 1993 were within a factor of 100 oflethal levels (48 h LC50 values expressed as concentra-tion factors ranged from <�10 to �96), although otherestuarine LC50 concentration factors lay in the range�100 to �730. Non-polar organic contaminants fromnearshore sites were also fairly close to lethal levels in

Table 1

Oyster embryo bioassay data for 1994 from UK estuarine and coastal

waters (after CEFAS, 1998)

Location Number of

stations

Percent net

response (PNR)a

Tweed estuary 3 �22 to 31

Blyth estuary/harbour 3 31 to 40

Tyne estuary 5 9 to 98

Offshore from Tyne 1 51

Wear estuary 4 �3 to 53

Tees estuary 7 14 to 100

Offshore from Thames 2 13 to 47

Central Channel 1 23

Tamar estuary 3 11 to 39

Offshore from Tamar 1 24

Mersey estuary 6 10 to 61

Dee estuary 2 5 to 39

All other offshore stations 10 �7 to 18

a A PNR value up to +20 is considered normal; a PNR of 21–49

indicates slightly impaired water quality; 50–99 indicates substantial

deterioration in water quality; and a PNR of 100 indicates very poor

water quality. Negative values indicate water quality better than the

reference water (obtained from the southwestern approaches of the

English Channel).


some cases (48 h LC50 values from �46 to �760). Onthe other hand, water extracts from offshore sites had tobe concentrated by 345 to >1000 times before causinglethality. The toxicity of estuarine and harbour waterextracts could be ranked in decreasing order of toxicityas follows: Tees>Wear>Mersey>Tyne, Blyth, PooleHarbour, Southampton Water>Ribble, Dee, Lune.Bearing in mind that the extraction technique does notpick up polar contaminants, these results imply that thewaters of industrialised estuaries and some coastal areasare sub-lethally or chronically toxic to some inverte-brates, thus supporting the observations made with theoyster embryo bioassay. Similar results have beenobtained in other European industrialised estuaries suchas the Rhine (Hendricks et al., 1994), but contaminantconcentration techniques have not yet been widely usedfor routine survey purposes.

3.2. Bioassays of sediments

A problem with ex situ bioassays of waters is that thesamples taken only represent an instant in time. Bycontrast, bioassay results obtained with sediments aremore likely to represent the long-term exposure situa-tion, so much attention has recently been given to thissubject in the UK (Thain et al., 1996; CEFAS, 1998). Inthe period 1990–1994, work focussed on the toxicity ofsediment elutriates, i.e. seawater shaken with surface (0–10 cm) sediment (200 ml sediment to 500 ml referenceseawater) for 3 h, and then filtered. These elutriateswere tested using the oyster embryo bioassay, whichmeasured toxicity (PNR values in the range 20–100) atseveral stations in six estuaries: Tweed, Tyne, Wear,

Tees, Thames and Mersey. Slight toxicity was alsodetected at two offshore sites in the English Channel,but high PNR values (50–100) were confined to a fewestuaries (Tyne, Tees, Thames and Mersey).These results suggested that some estuarine sediments

as well as waters might be acutely toxic, but the toxicityof elutriates was not considered representative of sedi-ments as a whole because it ignores the toxicity asso-ciated with ingested particles and with direct skincontact. Elutriation also tends to dilute the interstitialwater which is an important carrier of sedimentarycontaminants, and may artificially increase the bioa-vailability of some sediment-bound substances. Anextensive set of toxicity data (1992–1995) has thereforealso been obtained using whole sediment bioassays ofthe 0–10 cm layer with the polychaete lugworm Areni-cola marina and the amphipod crustacean Corophiumvolutator (CEFAS, 1998; Thain et al. 1996). In bothcases, test animals were exposed to sediment in thelaboratory for 10 days and mortality recorded, but inaddition, the rate of casting by A. marina (an indirectmeasure of feeding activity) was measured. Casting ratewas generally the most sensitive endpoint, and was sig-nificantly reduced at a large number of estuarine andharbour sites (typical data for 1992–1993 summarised inTable 2), including the Tweed, Tyne, Wear, Tees,Humber, Ouse, outer Thames, Southampton Water andPoole Harbour. Some effects on casting were also seenin a number of coastal and offshore sediments from theNorth Sea, English Channel and Irish Sea, althoughmost offshore sediments were not toxic. On the otherhand, significant mortality of A. marina was only seenin Wear and Tees estuary sediments, and significant

Table 2

Toxicity of whole sediment from estuarine and coastal areas to Arenicola marina and Corophium volutator in 1992–1993 (after CEFAS, 1998)

A. marina

% mortality

(no. of stations)

A. marina

% reduction

in casting

(no. of stations)

C. volutator

% mortality

(no. of stations)

Tweed estuary 0–7 (3) 0–35* (3) 13–30 (3)

Blyth harbour 7 (2) 7–19 (2) 13–23 (2)

Tyne estuary 0–10 (6) 0�96* (6) 7�40* (6)

Offshore from Tyne 0�13 (1) 13�27* (1) 10�17 (1)

Wear estuary 0�100 (7) 0�95* (7) 3�30* (7)

Offshore from Tees 0 (1) 0�16 (1) 0�17 (1)

Tees estuary 0�100 (5) 20�100* (5) 23�100* (5)

Humber estuary 0 (8) 15�59* (8) 0�30* (6)

Offshore from Humber/Wash 0 (1) 35* (1) –

Ouse estuary 0�20 (3) 0�57* (3) 0�3 (3)

Thames estuary (outer) 0 (2) 0�82* (2) –

Southampton Water 0 (2) 17�100* (2) 23 (1)

Poole Harbour 0 (1) 56* (1) 13 (2)

Offshore from Mersey 0 (1) 44* (1) 23 (1)

Mersey estuary 0�13 (4) 0�5 (4) 0�13 (3)

* Significant difference from reference sediment (P<0.05) in A. marina casting and C. volutator mortality data. >20% A. marina mortality is

considered adverse.


mortality of C. volutator only occurred in samples fromthe Tyne, Wear, Tees and Humber (Table 2). No coastalor offshore sediments were acutely toxic.More recent sediment bioassays of dredgings from

several UK industrialised estuaries and harbours (usingA. marina, C. volutator, and the amphipod Leptocheirusplumulosus in whole sediments, and the copepod Tisbebattagliai in pore waters) have confirmed that many ofthese sediments are acutely toxic to some sediment-dwellers (Reed, 1999). Included in these are dredgingsfrom Newport Docks, Swansea Docks, Blyth Harbour,and the Mersey, Tyne and Tees estuaries. Earlier work(Y. Allen, personal communication, 2000) had shownsimilar toxic effects of dredgings from five other UKdock and harbour areas (including Portsmouth, South-ampton, Falmouth, Cardiff and Hartlepool). Causativesubstances were not conclusively identified, but a num-ber of dredgings contained elevated levels of severalheavy metals and polychlorinated biphenyls (PCBs),and tributyltin (TBT) from antifouling paints was sus-pected of making a major contribution to toxicity at anumber of locations.There have been a few other more spatially limited

studies which also indicate that sediments from someindustrialised estuaries exert toxicity. One is by Lindleyet al. (1998) who showed that the hatching success ofcalanoid copepod eggs extracted from Mersey andHumber sediments was only 14 and 48%, respectively,compared with a figure of 92% for eggs from the Exe.These observations were closely linked to measuredconcentrations of various toxic substances. Matthiessenet al. (1998a) conducted a detailed survey of Tyne sedi-ments using a battery of eight bioassays, and found thatsediments from mid-estuary stations were generallymore toxic that those found in the lower or upper stret-ches. This pattern of peak toxicity also coincidedbroadly with the peak concentrations of a range ofsedimentary contaminants, but no single contaminantcould be clearly identified as a sole cause of toxicity. Itwas of particular note that the toxicity of apparentlyhomogeneous sediments varied markedly on a spatialscale of just a few metres, and it is possible that theapparently strong annual variation in sediment toxicityin some industrialised estuaries reported in CEFAS(1998) is at least partly an artefact caused by slightimprecision in the position of repeat sampling.Further studies have shown that metal-contaminated

sediments in the Fal estuary system have exerted effectson communities of nematodes. Millward and Grant(1995, 2000) have shown that nematode communities inRestronguet Creek, which have been exposed for over200 years to elevated metal loads from mining oper-ations, have responded by developing tolerance to cop-per toxicity, both in terms of increased abundance ofcopper-tolerant species, and evolution of enhancedtolerance in some species. This Pollution Induced

Community Tolerance (PICT) appears to be triggeredabove a threshold of effect of about 200 mg Cu/g sedi-ment. The fact that rapidly reproducing species andcommunities are able to develop tolerance to certaincontaminants is an important fact which should beborne in mind when assessing pollution effects in nat-ural sedimentary ecosystems (see below). Austen andSomerfield (1997) have also demonstrated using micro-cosms that metal-contaminated sediments from parts ofthe Fal system are able to alter community structure innematodes from a clean system. Furthermore, Schratz-berger et al. (2000b) have shown that dredgings fromthe Tees and Mersey adversely affect migration andsurvival rates of assemblages of nematodes fromuncontaminated sediments that have been exposed tosimulated dredgings disposal.

3.3. Effects of contaminants on water-column organisms

There are no reliable data which unequivocally linkcontaminants to effects on populations and commu-nities of UK marine water-column organisms, but areasonable body of information suggests that con-taminant-related impacts are occurring in individuals.Most of these have been detected in estuarine waters,but there are some reliable data on scope-for-growth(SFG) in mussels (Mytilus edulis) which show that con-taminants are also exerting effects in the coastal envir-onment. SFG is a measure of the amount of energyavailable to an organism for somatic growth, and thisvalue decreases in polluted organisms which are havingto devote energy to detoxification and tissue repair. Inextreme cases, SFG can be zero or negative, implyingthat the organism will fail to grow or actually loseweight and die. Data on SFG in mussels collected fromthe UK North Sea coast in 1990 (Widdows et al., 1995)show a fairly clear north–south trend, with high SFG inmost Scottish areas (with the exception of mussels col-lected in the vicinity of the Ythan estuary where a verylow SFG value of �1 J g�1 h�1 may have been causedby locally high levels of PCBs and/or algal toxins), andlower SFG as one moves south. From Blyth (nearNewcastle) southwards, mean SFG never exceeded 10 Jg�1 h�1, and from Harwich southwards, mean SFGremained below 5 J g�1 h�1. These observations com-pare with a normal value in uncontaminated mussels of>20 J g�1 h�1. Much of the effect could be explained interms of bioaccumulated polycyclic aromatic hydro-carbons (PAH), polar organics and tributyltin, butadditional unidentified toxicants were responsible for asignificant proportion of the effect in southern areas.More recently, a similar survey of SFG in mussels has

been conducted on the UK west coast from Scotland toCornwall (Widdows et al., 2002; EA, 1999). Thisrevealed generally higher SFG values than on the NorthSea coast, arranged in a different spatial pattern, with


high SFG values (10–25 J g�1 h�1) in the north andsouth, but medium values (5–10 J g�1 h�1) in the centralsection between the Lake District and North Wales.Once more, only a proportion of the total toxic effectcould be explained in terms of identified toxic sub-stances in mussel tissue, with PAH again forming thesingle most important chemical group. In broad terms,the SFG data show that UK coastal water quality in the1990s in areas where dense human populations dis-charge large quantities of domestic and industrial efflu-ent was less than optimal for normal mussel growth,and that no single group of contaminants was respon-sible for the whole effect. Recent work in Scotland(Gowland et al., 2000) has shown that mussels in thevicinity of an aluminium smelter in Loch Leven areexperiencing elevated activity of glutathione-S-transfer-ase in the hepatopancreas due to increased exposure toPAH, but no data are available on holistic biologicalmeasures such as SFG.Another effect in organisms from the open sea which

is less firmly linked to contaminants concerns fish larvalabnormalities. It has been well-established (e.g.Cameron et al., 1992; Cameron and von Westernhagen,1996; 1997) that malformation rates in the surface-floating embryos and larvae of many flatfish and otherfish species are very high (up to 30%) in UK and con-tinental coastal waters around the southern North Sea,while rates offshore are lower (9%). Along the Englisheast and north-east coast up to about 100 km offshore,abnormality rates in all stage Ia-II larvae combinedwere significantly in excess of the North Sea mean at upto six stations in 1991–1992. It has been shown experi-mentally that many of these abnormal larvae are des-tined to die. Although there is very high natural wastageof flatfish before they reach maturity due to predationetc., it seems possible that these unusual abnormalityrates could be adversely affecting populations.Cameron and her co-workers suggested that the

effects may be due to contaminants, but no proof of thishas been forthcoming, although recent circumstantialevidence indicates that contaminants are probablyinvolved (Von Westernhagen et al., 2000). Furthermore,recent work in the USA and elsewhere has shown thatultraviolet light (UV) is able greatly to enhance thetoxicity of PAHs and other organics through the pro-duction of highly destructive oxygen radicals in trans-lucent living tissues. As floating fish embryos areexposed to UV light, it has been suggested that theeffects seen in wild larvae may be due to an interactionof UV and the elevated levels of contaminants that havebeen measured at the sea surface. Experiments atCEFAS (B. Lyons, in press and in preparation) havetherefore been conducted to simulate the UV and PAHenvironment experienced by oyster embryos, and flatfishand cod larvae. These have shown, for example, thatalmost all sole (Solea solea) larvae or oyster (Crassostrea

gigas) embryos exposed for up to 72 h to environmen-tally realistic UV intensities can be killed or malformedby a benzo[a]pyrene concentration of 1 mg/l (and bysimilar concentrations of other PAHs such as fluor-anthene, anthracene and pyrene), while no effects occurin UV-free light. Such low levels of PAH had previouslybeen thought to be harmless to fish, but are known to bepresent in some places in the sea surface microlayer withwhich some pelagic fish embryos are in contact (Hardyet al., 1990; Kucklick and Bidleman, 1994; Zeng andVista, 1997). It remains to be seen whether such effectsare actually responsible for the field-observed phenom-ena, but the clear message is that one must be aware ofpossibly subtle interactions occurring between differentenvironmental stressors.As most of the major sewage and industrial discharges

are located on estuaries (or their inflowing rivers), it isnatural that much monitoring effort has been con-centrated in these areas. Particular attention has beengiven to the health of flounder (Platichthys flesus) andother flatfish using a variety of techniques to uncoversublethal responses to toxicants which may have long-term implications for fitness. The flounder is a commoneuryhaline flatfish which lives in intimate contact withsediments, and feeds on sedimentary invertebrates, so itis therefore considerably exposed to contaminants andcan be considered a useful sentinel species. The broadobjective has been to look for effects which are bothdiagnostic of causes and indicative of possibly moreserious impacts in the future.Early work with flounder in the Forth estuary (Elliott

et al., 1988) used the aryl hydrocarbon hydrolylase(AHH) assay to show that the activity of the cyto-chrome P450 detoxification system in liver was slightlyinduced in the area of Longannet which is contaminatedwith hydrocarbons. Subsequently, these observations inthe Forth were supported by Sulaiman et al. (1991)using the similar ethoxyresorufin-o-deethylase (EROD)assay (Burke and Mayer, 1974). This is now the stan-dard technique for measuring P4501A1 activity, andEROD induction in fish and other vertebrates can beconsidered a good biomarker of exposure to planarmolecules such as PAHs, polychlorinated biphenyls(PCBs) and polychlorinated dibenzofurans (PCDFs).It has since been shown that the cytochrome P450

enzyme system (as measured by EROD activity in liver)of plaice (Pleuronectes platessa) and dab (Limandalimanda) near the coasts of SW Wales was only slightlyand temporarily induced following the Sea Empress oilspill at Milford Haven in February 1996 (Kirby et al.,1999b), and other biological effects of the spill were alsotransient (e.g. Dyrynda et al., 2000; Fernley et al., 2000).On the other hand, the P450 system in dab from somestations in the vicinity of the earlier Braer oil spill in theShetland Islands in 1993 was more clearly induced forperiods of 6 months or more (Stagg et al., 1998). It


seems likely that these longer-term effects were due tohydrocarbons which had entered and persisted in sedi-ments during and after the very rough conditions pre-sent in the Braer incident.Perhaps of more long-term significance (due to on-

going exposure) are data from 1993/1994 on cyto-chrome P450 activity in pelagic fish larvae in the centralpart of the northern North Sea which show that ERODwas induced over wide areas and correlated reasonablywell with hydrocarbon concentrations in the water column(Stagg and McIntosh, 1996). Adult dab sampled fromthis area of intense hydrocarbon exploitation in 1990also showed elevated EROD activity near industrialinstallations (Stagg et al., 1995), as did juvenile plaicefrom the Firth of Forth in 1995 (NMP, 1998). Further-more, flounder sampled from several other industrialisedestuaries in 1997 (including stations on the Mersey,Tees, Wear, Tyne and Humber) had strongly elevatedEROD activity (Kirby et al., 1999b) in comparison witha clean reference site (Alde estuary). The highest activitywas observed in Mersey fish and was approximatelyfour times higher than in the Alde samples. EROD indab liver from the Thames estuary was also elevated in1991 in comparison with samples from the southernNorth Sea (Sleiderink et al., 1995).It is probable, though not conclusively established,

that both offshore and inshore effects described aboveare caused mainly by exposure to PAHs. This is sig-nificant because although the cytochrome P450 systemdegrades PAHs to more water soluble metaboliteswhich are rapidly excreted via the bile in fish, by-pro-ducts such as epoxides are genotoxic. The longer-termimplications of this are still being researched, but it hasalready been shown that levels of DNA adducts inflounder from the EROD-induced Tyne populationwere considerably elevated in 1997 (Lyons et al., 1999).This implies that metabolites produced by the cyto-chrome P450 system are indeed causing DNA damagewhich potentially (although not inevitably) can lead topre-neoplastic damage and even carcinoma in liver.Data are not yet available on disease rates in estuarineflounder, but it is known that EROD, pre-neoplasticliver nodules and actual hepatic neoplasias in anotherflatfish, the dab Limanda limanda, were all elevated in1993–1996 in certain offshore North Sea locations suchas the Flamborough area and Dogger Bank whichreceive contaminants from the northeast coast estuaries(CEFAS, 1998).Flounder from several UK estuaries are also experi-

encing a range of other sub-lethal stresses. One of theseis depressed acetyl- and butyryl-cholinesterase (ChE)activity in muscle (Kirby et al., 2000). Acet-ylcholinesterase is a vital enzyme which ensures thenormal degradation of the neurotransmitter acetylcho-line, and it is susceptible to inhibition by a range ofcarbamate and organophosphate pesticides. When ChE

is inhibited, nervous transmission is impaired, ulti-mately resulting in paralysis and death. Kirby et al.(2000) have shown that flounder sampled from severalstations on the Mersey, Tees, Humber, Tyne and Tamarestuaries in 1997 showed significant ChE inhibitioncompared with fish from the Alde, with the highestlevels of inhibition being present in Mersey and Tynefish (69 and 64% inhibition at Bromborough andScotswood Bridge, respectively). It is not yet knownwhether this enzyme inhibition is inducing behaviouraland other problems for flounder, or whether OP and Cpesticides are the sole cause. However, concentrationsof up to eight organophosphates and six carbamateswere above detection limits in all the surveyed estuariesexcept the Alde, so it is assumed that they were at leastcontributing to the observed effects.Studies have also been conducted in Scotland on

inshore plaice and flounder in the vicinity of a sewagesludge disposal site in the Firth of Clyde (Secombes etal., 1995), and on a mercury contamination gradient inthe Firth of Forth (Stagg et al., 1992). While it appearedthat the proximity of sewage sludge discharges in theClyde had little effect on various aspects of the immunesystem in plaice, there was an inverse relationshipbetween the concentration of mercury in flounder mus-cle and the degree of activity of branchial Na/KATPase, an enzyme which is vital for fish osmoregula-tion. Laboratory experiments have shown that mercuryis able to inhibit ATPase, but it is uncertain whether therather small effects seen in Forth flounder (approx. 35%reduction in enzyme activity to 6.5 mmol phosphate/mgprotein/h in the worst case at Longannet) were of sig-nificance for osmoregulation. A similar point can alsobe made about measurements of metallothionein activ-ity in Forth flounder (Sulaiman et al., 1991) whichshowed that this metal-binding protein was induced bya factor of up to 8 (to 270 mg/g liver) in fish from themetal-contaminated parts of the estuary.Perhaps of greater concern is the fact that many male

flounder in UK estuarine and coastal waters are show-ing signs of feminisation (Matthiessen et al., 1998b,2000; Allen et al., 1999a,b). This is manifested asinduction of the yolk protein precursor vitellogenin(VTG), often at concentrations in blood plasma severalorders of magnitude above reference fish (as high as 10million ng VTG/ml plasma, compared with 10–100 ng/ml in normal males). Estuaries in which VTG titres inmales and immature females were significantly higherthan at the Alde reference site during the period 1996–1999 included the Forth, Tyne, Wear, Tees, Humber,Crouch, Thames, Southampton Water, Mersey andClyde. Lower levels of VTG induction and occasionalovotestis have also been seen in male flounder makingtheir way offshore from contaminated estuaries to breed(Allen et al., 1999a), while normal VTG titres were pre-sent in male flounder from the Tamar and Welsh Dee.


Elevated VTG titres were often accompanied byincreased liver size as the liver is the site of VTG synth-esis and experiences hypertrophy and hyperplasia inresponse to oestrogenic stimuli. Furthermore, male fishfrom the Mersey and Tyne estuaries contained elevatedoestradiol titres compared with negligible titres in malescaught in the North Sea (Scott et al., 1999), and up to18% of males from these estuarine locations are intersexindividuals exhibiting ovotestis. Such fish have primaryand sometimes secondary oocytes in their testes and areprobably not able to produce sperm in normal quantityor quality. Research with other species of fish (Para-lichthys dentatus and Oncorhynchus mykiss) has shownthat such high levels of VTG induction in males are alsoassociated with disruption of spermatogenesis and renalglomerular damage (Folmar et al., 2001; Herman andKincaid, 1988), although these types of pathology havenot been fully studied in UK flounder. Gonadosomaticindex and phenotypic sex ratio of UK flounder haveboth been recorded but no obvious effects seen in oes-trogen-contaminated estuaries.The full significance of these effects was investigated

within the EDMAR programme (EDMAR, 1999),especially the implications for population survival.Recent work (Matthiessen et al., 2000) has shown thatmale viviparous blenny (Zoarces viviparus) from theClyde and Tyne are also synthesising VTG and showingsigns of ovotestis. It is hoped to link these effects toreproductive success because it is easy to assess numbersand quality of fry in these viviparous fish. Indeed, it hasalready been shown that the sex ratio of offspring in Z.viviparus has been shifted towards males in fish livingnear a pulpmill discharge in Sweden (Larsson et al.,2000). Furthermore, the males of two species of sandgoby (Pomatoschistus minutus and P. lozanoi) from theTees and Mersey are showing high prevalences (up to75%) of urogenital papillae which are intermediate instructure between male and female. These papillae aresecondary sexual characteristics that probably developunder hormonal control, so it seems likely that thefeminised papillae are related to oestrogen exposure.Indeed, recent experiments have shown that oestradiol-exposed juvenile male P. minutus develop with feminisedpapillae (M. Kirby, personal communication, 2001).The precise causes of all these feminising effects are notyet fully known, but there is evidence that both syn-thetic oestrogen mimics (e.g. nonylphenol) and naturaloestrogenic hormones (e.g. 17b-oestradiol) are con-tributing (see section below on cause–effect relation-ships; Thomas et al., 2000, 2001).

3.4. Effects of contaminants on benthic communities

The evidence, either recent or otherwise, for con-taminant impacts on UK benthic marine communitiesis not abundant. Indeed, in coastal and offshore

macrobenthos assemblages, there are no obvious effectsof the relatively low levels of metallic contaminants onbiodiversity, and no other indications of changes whichcannot be attributed to anthropogenic activities such astrawling, or natural variables such as sediment type andstability (CEFAS, 1998; Rees et al., 1999). Local excep-tions concern the effects of sewage sludge disposal atlicensed offshore sites (halted in 1998; CEFAS, 1997),and impacts of the continuing disposal of dredgedmaterials (also at licensed offshore sites; CEFAS, 1998).It is likely that the impacts of these activities on benthiccommunities are largely caused by organic enrichmentand smothering, respectively, although toxic effectscannot be excluded in all cases. However, there is alsogood evidence that offshore discharges of oil-baseddrilling fluids by the oil and gas industry (halted in1996) have caused reductions in benthic species diversityor other changes in community structure at distances of<1–3 km from drilling locations (Davies et al., 1984;Gray et al., 1990; Olsgard and Gray, 1995; Gray et al.,1999).Turning to estuaries, although there is much evidence

that estuarine benthic organisms are heavily exposed,for example, to metals (Bryan and Langston, 1992),there are few documented cases of consequent effects oncommunities, such as at Restronguet Creek in the Falestuary system where copper and zinc pollution frommining is strongly suspected to have caused the exclu-sion or restriction of several species of bivalves, includ-ing Cerastoderma edule, Macoma balthica, Mytilusedulis and Scrobicularia plana (Bryan and Gibbs, 1983),as well as the changes in nematode metal tolerancedescribed above (Somerfield et al., 1994). This generallack of observed effects is partly due to the fact thatsome species such as Nereis diversicolor are able toadapt to pollution (e.g. Grant et al., 1989), but perhapsmainly because of the difficulty of attributing causes toparticular changes in benthic communities. However,recent studies of benthic meiofauna (Schratzberger etal., 2000a) around the English and Welsh coasts haveshown that a combination of particle size and con-centrations of chromium, zinc, arsenic, mercury andcadmium best explain nematode and harpacticoidcopepod assemblage structures at inshore locations.Meiofaunal abundance and diversity were generallylower inshore than offshore, and in at least one inshorelocation (Burbo Bight in Liverpool Bay), high con-centrations of heavy metals mainly derived from pol-luting inputs were correlated with low diversity.A good illustration of the difficulty of demonstrating

pollution impacts on benthic communities concerns theTees estuary, which has received large volumes ofuntreated or poorly treated industrial effluent since the1920s, and much untreated domestic effluent sincethe early nineteenth century. Despite the grossly pol-luted nature of this estuary, a detailed study in 1971–


1973 (Gray, 1976) of the benthic invertebrate commu-nities in the outer reaches (Seal Sands and Gare Sands)failed to reveal clear evidence of impacts on the meio-fauna, although there had been some reductions inmacrofaunal species richness (especially polychaetes andbivalve molluscs) since a qualitative survey conducted inthe 1930s (Alexander et al., 1935). The physical natureof the estuary had changed over this period due to landreclamation and dredging, so it was unclear to whatextent the faunal changes were due to pollution. How-ever, partial clean up of discharges to the estuary beganin 1970, and over the period 1979–1985 there was astrong recovery in macrofaunal species richness (froma total of 33 species in 1979 to 78 species in 1985), with asimultaneous increase in numbers of individuals (Shil-labeer and Tapp, 1989). There have also been increasesin species diversity and spatial extent of macroalgae inthe Tees over a similar period (Hardy et al., 1993).There has been little improvement in Tees invertebratecommunities in more recent years (Tapp et al., 1993),and some polluted parts of the estuary still had animpoverished benthic fauna in 1990. These studies showthat even in heavily polluted locations, it may be neces-sary to make studies of temporal trends in benthiccommunities before and after management actions inorder to establish cause–effect relationships. Despite therelative abundance of data from the Tees, the precisecauses of benthic community impoverishment areunclear, although high biological oxygen demand andcontinuing industrial pollution have undoubtedlyplayed a part, as have natural environmental events inthe North Sea (Warwick et al., 2002).Other recent examples (1985–1999) of relatively

impoverished benthic invertebrate communities havebeen documented for parts of the Forth, Tyne, Humber,Thames, Southampton Water, Mersey and Clyde estu-aries (Moore, 1987; McLusky, 1987; Warwick et al.,1987; Oyenekan, 1988, McLusky et al., 1993; EA, 1999).These effects have been mainly attributed to organicenrichment from sewage discharges, although industrialpollution in some locations (e.g. Mersey) may also be afactor. It should be noted that in many cases, significantrecoveries in benthic communities have occurred overthis period as a result of continuing improvements in theclean-up of sewage and other discharges.There is, however, one clear example of pollution-

induced change in benthic ecosystems that is firmlylinked to a particular toxic pollutant. This concerns theantifouling paint ingredient tributyltin (TBT), theendocrine disrupting properties of which have beenshown to cause imposex in gastropods (Harding et al.1999) and to affect not just coastal and estuarine mol-lusc populations (Bryan and Gibbs, 1991; Matthiessenand Gibbs, 1998), but also cause reductions in speciesdiversity in estuarine benthic and epibenthic inverte-brate communities. In the Crouch estuary, where fresh

TBT inputs were largely halted in 1987 when the use ofTBT-based antifoulants was banned on small vessels,total numbers of benthic invertebrate species increasedfrom 37 in 1987 to 63 in 1991, with the largest recoveriesoccurring in the upper estuary areas where TBT con-tamination had been at its worst. Similar recoverieswere seen in the epibenthic fauna. It is of interest thatalthough molluscs made a strong comeback (not unex-pectedly, due to TBTs known molluscididal properties),many crustaceans and ascidians also re-appeared, indi-cating that the benthic and epibenthic communities as awhole had been severely damaged in the 1980s (Rees etal., 1999; Waldock et al., 1999). TBT inputs in manyestuarine areas reduced considerably during the 1990s,however, and community-level effects which weremarked in the late 1980s and early 1990s can no longerbe detected in the Crouch (Rees et al., 2000).Overall, it is apparent that while community-level

changes are the ‘bottom line’ when assessing the severityof pollution effects, they have little potential to bediagnostic of causes, and one-off surveys are very unli-kely to show clear cut effects which can confidently beascribed to particular pollutants. This may partiallyexplain the apparent mismatch between the detection ofwidespread estuarine sediment toxicity, on the onehand, and the relatively few reports of benthic commu-nity damage attributed to toxicants, on the other. It is,however, also possible that the mismatch is caused inpart by the adaptation of estuarine populations to pol-lution, a factor which would not be detected in sedimentbioassays which employ test organisms without a sig-nificant pollution history.

4. Attribution of cause–effect relationships for con-

taminants and organisms

Data obtained by Kirby et al. (1998) showed that asignificant group of contaminants in toxic hexaneextracts of seawater from estuarine and coastal stationswas total hydrocarbons (THC). There was a significantthough weak negative correlation (between �0.36 and�0.43) between THC and the 48 h LC50 concentrationfactor (based on acute toxicity to T. battagliai), withhighest THC concentrations found in the waters ofindustrialised estuaries (0.8–48 mg/l). At five of thesesites, total polycyclic aromatic hydrocarbons (PAH)were estimated to be contributing between 5 and 18% oftotal toxicity. However, it was clear that the majority ofthe toxic material had not been identified, and otherwork (e.g. Matthiessen et al., 1993) had shown thatwhile some UK estuaries contain large numbers (>70)of potentially toxic organic chemicals in the mid- tonon-polar range, hardly any of these are present atindividually harmful concentrations. The major excep-tions to this appear to be the polycyclic aromatic


hydrocarbons (PAH), some of which occur in UKestuarine waters and sediments at acutely or chronicallytoxic levels (Law et al., 1997; Woodhead et al., 1999),and whose toxicity can be potentiated by ambient levelsof ultraviolet radiation (B. Lyons, in press and in pre-paration.). It will nevertheless be apparent from theforegoing that clear identification of causes has rarelybeen achieved in field studies, and most investigatorshave at best relied on the derivation of correlationsbetween effects and their presumed causative con-taminants.The best way of tackling the question of causality in

such a complex situation is to use Toxicity Identificationand Evaluation (TIE) techniques which were firstdeveloped by the US Environmental Protection Agency(Mount and Anderson-Carnahan, 1988). In essence,these employ a variety of procedures to fractionatecomplex samples, each fraction of which is then bioas-sayed to identify the fraction or fractions containing thetoxic material. The toxic fractions are then further frac-tionated and bioassayed until sufficiently simple frac-tions have been produced to allow the toxic substance(s)to be identified by traditional chemical methods.Recent TIE investigations of UK estuarine and

coastal waters using acute toxicity to the bioassayorganism T. battagliai as their endpoint (Thomas et al.,1999a,b) have shown that a proportion of the effects inthis organism can be ascribed to alkylphenols, variouschlorophenols, alkyl-substituted naphthalenes, alkyl-substituted fluorenes, atrazine and dimethyl benzoqui-none. However, other causative substances remain to beidentified, and there is a range of suspects includingphthalates, triphenylphosphine sulphide, methyl acri-dine and dieldrin. The particular mix of causative sub-stances seems to be largely estuary-specific (Table 3).Similar methods have been used to identify the oes-

trogenic activity in the Tyne and Tees estuaries (Thomaset al., 2000, 2001). In these cases, the bioassay organismwas a yeast cell line that had been stably transfectedwith the human oestrogen receptor (ERa) gene linked toa reporter gene system. Waters in the vicinity of theHowdon and Dabholm Gut STW discharges werefound to contain the following oestrogenic substances:17b-oestradiol, androsterone (an androgen with oestro-genic activity), bis(2-ethylhexyl) phthalate, and non-ylphenol. The majority of the activity in water wasattributable to 17b-oestradiol (E2), but total activity atits most concentrated only amounted to 24 ng E2 equiv./l, insufficient to explain the high levels of feminisation inwild flounder caught locally. However, much higheramounts of oestrogenic activity (up to 4.7 mg E2 equiv./kg)were observed in sediments, mainly associated with thesolid phase. Most of this activity remains to be identified,although a proportion is attributable to nonylphenol.It should be borne in mind that definitive and com-

prehensive results from TIE procedures cannot be

obtained through the use of a single bioassay, as differ-ent organisms will be susceptible to different substances.However, although the use of TIE techniques in UKwaters is still in its infancy, the method clearly haspotential for pinning down the causes of toxic effects.

5. Conclusions and recommendations

Although the spatial distribution of a fairly limitedsuite of anthropogenic substances in the waters, sedi-ments and some biota has now been established formany UK estuaries and coastal waters, it is difficult toassess the biological significance of much of this infor-mation because of mixture effects and the poorlyunderstood issue of bioavailability. Furthermore, fewreliable data exist on the temporal trends of most con-taminants, although there is no doubt that the dis-charged loads of some substances are decreasing. Forthese reasons, there is a need to assess the biologicaleffects of chemical contamination in a more direct way.Table 4 represents a summary of some UK estuarine

biological effects data for the period 1985–1999, focus-ing on those estuaries and pollution responses for whicha reasonable range of data are available. This gives arather crude picture because it does not show the partial

Table 3

Substances present in UK estuarine waters which have been tentatively

identified by TIE as contributing to toxic effects in the copepod Tisbe

battagliaia

Estuary Substance

Tyne Chlorobenzene acetonitrile

4-Chloro-3,5-xylenol

Methyl acridine

Methyl pyridine amine

Monomethyl-trimethyl fluorenes

Monomethyl-trimethyl naphthalenes

Naphthalene amine

Nonylphenol

Pentachlorophenol

Trichlorophenol

Tetrachlorophenol

Triphenylphosphine sulphide

Tees Atrazine

Carbophenothion methyl sulphoxide

4-Chloro-3,5-dimethylphenol

Diethyl naphthalene carboxamide

Dimethyl benzoquinone

Dimethyl naphthalene carboxamide

Nonylphenol

Mersey Dieldrin

Dodecylphenol

Milford Haven Bis(2-ethylhexyl)phthalate

a Other contributary substances remain to be identified. The data

are based on Thomas et al. (1999a,b).


recoveries that have been made over the period in anumber of estuaries subject to discharge clean-up pro-grammes. Furthermore, most available datasets ofbioassay and biomarker responses are insufficientlydetailed to reveal whether temporal trends are all in thedirection of improvement, although most contaminantdatasets suggest that this is the case. However, Table 4does allow a rough ranking of the breadth of biologicalimpact, as follows (in order of decreasing numbers ofimpacts):

Mersey>Tyne=Tees>Humber>Thames>Forth>Wear>Clyde>Southampton Water>Crouch>Tamar>WelshDee

This ranking of course excludes a large number ofestuaries for which few data are available, and alsoexcludes responses for which there are no data in aparticular estuary. It should not therefore be inferredthat the biological quality of the Mersey and Welsh Deeare necessarily the worst and best, respectively, to befound in the UK. There are many which receive verylittle in the way of point source inputs and in which theeffects of pollutants are likely to be negligible. One suchexample is the Alde in Suffolk where plasma vitellogenintitres in male flounder in 1996–1997 (a measure of theamount of oestrogenic exposure derived largely fromsewage effluent) were actually lower than in fish caughton their spawning grounds in the North Sea (Allen etal., 1999a). Indeed, it is apparent from this review that

UK coastal and offshore waters and sediments cannotbe considered entirely unimpacted by contaminants,although in general they are likely to be less affectedthan most estuaries. In support of this statement, thereis no evidence of significant pollution effects on benthiccommunity structure on the continental shelf (outsidethe immediate vicinity of oil production installationsand dredgings disposal grounds; Rees et al., 1999).Nevertheless, a number of biomarkers indicate thatindividual organisms in some open sea locations are atleast being exposed to contaminant levels capable ofprovoking a response at the biochemical, cellular orphysiological levels (e.g. EROD induction in pelagic fishembryos in the northern North Sea—Stagg and McIn-tosh, 1996; increased prevalence of hepatic neoplasia insome North Sea dab—CEFAS, 1998; depression ofscope-for-growth in coastal mussels—Widdows et al.1995, 2002).This raises the issue of how to interpret the true bio-

logical significance of biomarker changes. It is clear thata change at the level of an individual (including itsdeath) does not necessarily imply that the populationand community of which it is a member will be sig-nificantly affected. It is a difficult theoretical problem toextrapolate with confidence from the individual tohigher organisational levels (for a useful discussion, seeCalow and Forbes, 1998) due to such factors as density-dependence which can lead, for example, to increasedgrowth and reproductive rates in survivors if someindividuals in a population are eliminated. Our present

Table 4

United Kingdom estuaries in which the biological effects of pollutants have been detected during the period 1985–1999

Forth Tyne Wear Tees Humber Crouch Thames So’ton

Water

Tamar Welsh

Dee

Mersey Clyde

Benthic community impoverishmenta Yes Yes Yes Yes Yes Yes Yes Yes No No Yes Yes

Crassostrea oyster embryo water

bioassay responsebYes Yes Yes Yes Yes No Yes No Yes Yes Yes No

Tisbe bioassay of water extracts

(48h LC50 conc. factor4100)cn.d. No Yes Yes n.d. n.d. n.d. No No No Yes n.d.

Arenicola and Corophium sediment

bioassay responsedn.d. Yes Yes Yes Yes No Yes Yes No No Yes n.d.

Vitellogenin induction in male fishe Yes Yes Yes Yes Yes Yes Yes Yes No No Yes Yes

Intersex in male fishf No Yes No No No No No No No No Yes No

EROD induction in flatfishg Yes Yes No Yes Yes n.d. Yes No n.d. n.d. Yes n.d.

ChE inhibition in flatfishh n.d. Yes No Yes Yes n.d. n.d. n.d. Yes n.d. Yes n.d.

n.d., No data.a Moore (1987); McLusky (1987); Warwick et al. (1987); Oyenekan (1988); McLusky et al. (1993); NMP (1998); EA (1999); Rees et al. (1999,

2000); Waldock et al. (1999). It should be noted that in many cases the benthic fauna have been impoverished mainly as a result of organic enrich-

ment, rather than overt toxic effects. Furthermore, many of the reported effects were declining by the mid-1980s, and continuing improvements in

sewage treatment have in several cases led to further recoveries in benthic fauna since then.b NMP (1998); CEFAS (1998); J.E. Thain personal communication (2001).c CEFAS (1998); Kirby et al. (1998).d CEFAS (1998).e Matthiessen et al. (1998b, 2000); Allen et al. (1999a,b).f Matthiessen et al. (1998b, 2000); Allen et al. (1999a,b).g Sleiderink et al. (1995); CEFAS (1998); NMP (1998); Kirby et al. (1999a).h Kirby et al. (2000).


level of understanding does not permit a reliable ecolo-gical interpretation of biomarker changes, so we have tomake do with qualitative or weight-of-evidence assess-ments of the available data.Such an assessment of the data summarised above

indicates that there is a group of highly industrialisedestuaries (Mersey, Tyne, Tees, Humber, Forth andThames) where most indicators show that toxicants arehaving biological effects at the level of the individual inrepresentatives of molluscs, crustaceans and fish. Theseare accompanied by effects at the benthic communitylevel, although it is probable that most of these are attri-butable to organic enrichment rather than toxicity. Thereis another group of estuaries exposed to less intense orlargely historical pollution sources (Clyde, Wear, South-ampton Water and Crouch) where less than 50% of bio-logical pollution indicators give a response, whileestuaries giving or expected to give a response in less than30% of indicators (e.g. Tamar and Dee) are almost cer-tainly in a majority in the UK as a whole. It is perhapsnoteworthy that the Tamar and Dee are the only estu-aries listed in Table 4 where benthic community effectsattributable to pollution do not appear to be occurring.In summary, it is clear that a wide range of pollution-

related biological impacts is still present in the UK’smore heavily urbanised estuaries, some of which areprobably having knock-on effects on populations andcommunities. On the positive side, however, most con-taminants for which we have data on temporal trendsare declining in significance, although we need toremain vigilant in the detection of new groups of con-taminants such as polybrominated diphenylethers. Inview of the complex mixtures of contaminants presentin UK estuaries (Matthiessen et al., 1993), it will beinsufficient to improve our environmental risk assess-ment programmes for chemicals without also imple-menting Direct Toxicity Assessment procedures onindividual discharges, and continuing chemical andbiological monitoring of receiving waters (Matthiessen,1998).A number of recommendations flow from these studies:

1. Efforts to clean up point-source discharges anddiffuse inputs to the UK marine environmentshould continue, as serious biological effects ofcontaminants are still occurring in some urba-nised estuaries and coastal waters

2. Due to the problem of mixtures, aquatic envir-onmental risk assessment of chemicals and reg-ulation by Environmental Quality Standards areboth inherently imperfect procedures which mustbe supplemented with toxicity assessment of dis-charges, receiving waters and sediments

3. It is important to be alert for interactions ofapparently unrelated environmental stressorssuch as PAH toxicity and ultra-violet radiation

4. Efficient monitoring of the marine environmentcan only proceed if chemical and biological sur-vey techniques are fully integrated

5. In order to assist managers of the marine envir-onment, it is important to follow up positivebiomarker and bioassay responses with ToxicityIdentification and Evaluation procedures whichare able to pinpoint the causative substances

6. For collaborative marine monitoring pro-grammes, especially those involving measure-ments over time, increased effort is required toensure continuity in sampling design and con-sistency in data quality

7. More biomarkers that are themselves diagnosticof causes need to be developed in support ofmarine monitoring programmes

8. Further basic research is required to allow amore confident ecological interpretation of thesignificance of biochemical, cellular and physi-ological responses to pollution

Acknowledgements

The results described in this paper have been pro-duced by many scientists, especially those at the Centrefor Environment, Fisheries and Aquaculture Science(CEFAS). Particular thanks for their scientific con-tributions are due to the following CEFAS staff: ColinAllchin, Yvonne Allen, Steve Feist, Mark Kirby, BrettLyons, Steve Morris, Hubert Rees, Sandy Scott, JohnThain, Kevin Thomas and Mike Waldock. Most of theCEFAS work described in this paper was funded by theUK Ministry of Agriculture, Fisheries and Food(MAFF—now DEFRA), the UK Department ofEnvironment, Transport, and the Regions (DETR—now DEFRA) and the England and Wales EnvironmentAgency (EA).

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