BIOMONITORING OF TRACE METALS IN ESTUARINE AND MARINE ENVIRONMENTS

Transcription

1 BIOMONITORING OF TRACE METALS IN ESTUARINE AND MARINE ENVIRONMENTS PS Department of Zoology, The Natural History Museum, Cromwell Rd, London SW7 5BD, UK. Manuscript received, 30/9/2006; accepted, 6/12/2006. ABSTRACT In assessing the distributions of trace metals in aquatic environments as a prerequisite to any study of the variation of local toxic metal contamination over space and time, it is tempting to measure dissolved and/or sediment concentrations. A third and preferred alternative, however, is the measurement of accumulated trace metal concentrations in selected organisms, termed biomonitors. Such concentrations are high enough to be easily measured without significant risk of contamination. Most significantly, the accumulated concentrations are time-integrated relative measures of the amounts of metal that have been taken up into the organism via all uptake routes, by definition the bioavailable metal - the ecologically significant ambient metal in a habitat given its potential for ecotoxicological effects. To be selected as a biomonitor, an aquatic organism must fulfil several characteristics, and importantly we need an understanding of its accumulation kinetics for a particular metal, as might be described for example by biodynamic modelling. It is preferable to use a suite of biomonitors to investigate the several possible metal sources (solution, suspended material, newly deposited or older sedimented material, etc.) in an aquatic habitat. The choice of this suite should ideally include cosmopolitan biomonitors for which a database of accumulated metal concentrations has been compiled, therefore allowing the identification of trace metal availabilities considered high on an international scale. A major caveat is the need to avoid interspecific comparisons of accumulated metal concentrations, given the wide variation in accumulation patterns between even relatively closely related taxa. Case studies are used to illustrate the principles of the biomonitoring of trace metals in estuarine and marine environments. Threshold accumulated trace metal concentrations in key biomonitors have the potential to be used as simple measures identifying the presence of ecotoxicologically significant trace metal pollution in a coastal habitat. Key words: trace metals; bioavailability; biomonitors; biodynamic modelling. INTRODUCTION This paper addresses one aspect of the ecotoxicology of trace metals in aquatic environments crudely, how to answer the question How does toxic metal pollution vary over space and time in and between aquatic habitats?. The title is deceptively simple, yet even a term like trace metals is by no means straightforward to define. Biologists, including ecotoxicologists, tend to use the term without strict definition, sometimes synonymously with another term heavy metals. The term trace metal suggests occurrence at low (trace) concentrations (sometimes defined as 0.01% dry weight) in the environment, in both physical and biotic components, yet all of these metals do occur in raised concentrations in ores and some spectacularly so in organisms, as will be demonstrated. An implication of the label heavy metals is that these are metals above a threshold atomic weight, typically incorporating all transition metals of the periodic table. Yet the larger metals of the first two groups of the periodic table (for example Group 1 alkali metals like caesium or Group 2 alkaline earth metals like barium or radium), and the actinides and lanthanides, are usually not considered to be heavy metals, although their weights exceed those of some transition metals. So a characteristic other than weight is implicit in fact a chemical criterion. Thus these metals of interest show similar chemical characteristics indeed the characteristics that that make them biologically relevant yet to differing degrees according to their positions in the periodic table, thereby blurring even any chemical definition. These metals are toxic to biota when present in high bioavailability and yet many are essential to the metabolism of life, consistently across the eukaryotes with excellent but not perfect agreement between eukaryotes and prokaryotes. Many of the biochemical pathways underlying life processes are conserved in all organisms and require the same elements to function. Indeed some authors, particularly nutritionists, restrict the term trace metal to essential metals and exclude the non-essential metals those that have not (as yet) been shown to perform any essential metabolic function in organisms. In this paper, the term trace metal will be defined very widely and will include both essential and non-essential metals. For the chemists, a strict definition would be that of Nieboer and Richardson (1980) who proposed a chemical classification system based on the Lewis acid properties of metal ions, separating these into Class A, Class B or Borderline according to their degree of hardness or softness as acids and bases. Class A metal ions are Lewis hard acids, readily form cations (are more ionic ) and have a ligand affinity order O > N > S, while Class B metal ions are Lewis soft acids, are more covalent and have an affinity order S > N > O. Borderline metal ions have intermediate properties. Metals with Class B or Borderline ions fit into the category of trace metals as used here. For the non-chemists, Table 1 is a non-comprehensive list of the elements loosely defined as trace metals for the 107

2 Table 1. A non-comprehensive list of trace metals as defined for the purpose of this paper. purpose of this paper, including elements like arsenic and selenium commonly referred to as metalloids. It is the affinity of trace metals for sulphur and nitrogen that promotes their binding to molecules in cells, particularly to proteins, and that makes some of them essential and all of them toxic, binding in the wrong place at the wrong time when available in excess. That is why they are of interest to ecotoxicologists. In addressing the question How does toxic metal pollution vary over space and time in and between aquatic habitats?, we are not yet addressing the and so what? question. Whether increased levels of trace metal pollution (strictly, raised bioavailabilities of trace metals to biota) are significant enough to cause ecotoxicological effects is the key follow up question for an ecotoxicologist, and we now have a battery of techniques to address this question based on biomarkers. A biomarker is a biological response (for example a biochemical, cellular, physiological or behavioural variation) that can be measured in tissue or body fluid samples or at the level of the whole organism, that provides a measure of exposure to and/or effects of a contaminant (Depledge 1989; Peakall 1992; Depledge and Fossi 1994; Luoma and in press). A pre-requisite to the so what? question is to know the distribution in time and space of the highest toxic metal bioavailabilities, so that we can then assess their ecotoxicological significance and also identify their sources. This paper will show the key role of trace metal biomonitors in this procedure, defining a biomonitor as an organism which accumulates trace metals in its tissues, the accumulated metal concentration of which may be analysed to provide a relative measure of the total amount of metal taken up by all routes by that organism, integrated over a preceding time period (Luoma and in press). MEASUREMENTS OF TRACE METAL POLLUTION There are three major categories of trace metal concentration data measured in attempts to compare differences in trace metal pollution in aquatic habitats over space and/or over time trace metal concentrations in water, in sediment and more recently, in resident biota. Water When deciding how to measure trace metal pollution in aquatic habitats, it is logical to consider the measurement of dissolved trace metal concentrations. After all, there is a vast database in the literature defining the dissolved concentrations of trace metals that have proved toxic to a range of biota in toxicity tests. The technology now exists to measure extremely low dissolved concentrations of trace metals. Nevertheless, in many aquatic habitats, dissolved concentrations are still close to the limits of detectability and, more importantly, nearly all measurements are at risk from contamination during collection and analysis. Extreme precautions need to be taken with appropriate sampling and clean laboratory techniques to produce reliable concentration results. Furthermore, dissolved concentrations usually vary over time, particularly for example in estuaries with differential inputs of river and sea water at different states of the tide, and differential river flow according to recent rainfall in the catchment often varying seasonally. Each measurement represents a single time point that may be very different from the dissolved concentration present at that exact location the day before or the day after. It is thus necessary to design extensive sampling programmes over time that can account for such variation and give a more realistic picture of the amount of dissolved metal present in a habitat. Such sampling programmes are achievable but are expensive. It is also important to ask what such data mean, however accurately they have been measured. The data are measures of total concentrations dissolved in the water (usually above a defined cut-off size typically 0.45 µm). However, the physicochemistry of the medium will affect the rate of uptake of a metal from solution by biota. In short, the bioavailabilities of metals in different water bodies may vary even when their total dissolved concentrations are identical. For example the uptake rates of zinc and cadmium by biota from solution typically increase with decreased salinity over the salinity range found in an estuary even if the dissolved concentrations are unchanged ( 1995a). Thus comparisons of dissolved concentrations between habitats may lead to misleading conclusions on the relative bioavailabilities of dissolved toxic metals in the habitats (Luoma and in press). Therefore, accurate dissolved concentrations require extensive and expensive monitoring programmes. The data obtained are measures of total, not bioavailable, metal concentrations, and would need considerable further manipulation, for example by using physicochemical speciation modelling, to obtain more accurate estimates of relative dissolved bioavailabilities. It is the latter that are of interest, for they alone represent 108

3 measures of the ecologically significant ambient dissolved metal in the habitat with the potential for ecotoxicological effects. It does remain true, of course, that gross differences in total dissolved metal concentrations will correlate with differences in dissolved metal bioavailabilities. This is not as reassuring as it first appears, because dissolved concentrations in contaminated aquatic habitats showing ecotoxicological effects are usually below dissolved concentrations used in toxicity experiments, and the diet of animals cannot be ignored as a potential route of ecotoxicologically significant metal input in the field (Wang 2002; Luoma and 2005). This dietary uptake route will supply additional metal of ecotoxicological potential, not reflected in estimates of dissolved bioavailabilities alone. Furthermore, dissolved trace metals will typically partition out into the sediments, and large differences in anthropogenic input of metals into aquatic habitats (resulting, for example, from differential mining activities) are not always well reflected in the resulting changes in dissolved metal concentrations in the receiving habitat. Comparisons of expensively-gained dissolved metal concentrations will therefore often not supply the comparative information on the ecotoxicological potential of metals in different aquatic habitats that their cost would imply. SEDIMENT The second common measure of trace metals in aquatic habitats is their concentration in the local sediments. Metal concentrations in aquatic sediments have advantages. They are typically high enough to be easily measured without danger of significant contamination during collection and analysis. Sediments also show some integration of accumulated metal over time. Thus the concentration of metal in the sediment represents metal that has been accumulated over a period of time, and any measurement made does not have the same potential as dissolved concentrations to vary over short time periods. Thus, there is not the same need for extensive sampling programmes as would be required for dissolved concentrations. There are still, however, problems with the use of sediment metal concentrations as comparative measures of trace metal pollution. The physicochemical characteristics of sediments greatly affect their ability to accumulate trace metals, so that different sediments will reach different concentrations from identical dissolved sources of metal (Luoma 1989, 1990; Bryan and Langston 1992). For example particle size and organic carbon content, usually inversely covarying, have major effects on metal accumulation by the sediment. Sediments high in organic carbon content will bind more metals than those with low organic carbon, and sediments with particles of high surface area will also accumulate more metals than particles of low surface area. Thus muds (small particle size, high organic content) accumulate more trace metals than sands (large particle size, low organic content). In the mixing zones of estuaries, river-borne iron and manganese will precipitate from solution as hydroxides as a result of increasing ph. As oxides and hydroxides of iron sink to be incorporated into the sediment, many trace metals such as Ag, As, Cu and Pb (Bryan and Gibbs 1983) adsorb onto them. Correspondingly, differences in the iron oxide contents of aquatic sediments will cause differences in the capacity of the sediments to accumulate trace metals from solution (Luoma and Bryan 1978, 1982, Luoma 1989, 1990, Bryan and Langston 1992). Sediments can also be a source of trophic metals to sedimentingesting animals and the trophic availability of metals in ingested sediments will also vary with the physicochemical characteristics of the sediment (Luoma and in press). Again, therefore, total metal concentrations in sediments may not be good measures of the relative bioavailabilities of trace metals in different sediments, in this case as a source of trophically available metal for biota. It is also possible that sediments in the different aquatic habitats under comparison have been derived from the erosion of rocks with different mineral composition, including constitutive trace metal contents. Such constitutive metal contents will contribute to total metal concentrations measured (according to the nature of extraction or digestion used in the preparation of the sediment for analysis), and detract from their accuracy as comparative measures of the more recent (anthropogenic?) supply of trace metals. Total metal concentrations of metals in sediments are, therefore, not necessarily accurate relative measurements of metal bioavailabilities in compared habitats. In an attempt to overcome this problem, sequential extraction procedures may be used (Tessier and Campbell 1987). Accumulated metals are extracted with sequentially stronger chemical extraction, and ultimately digestion, procedures in an attempt to model what might be considered to be the bioavailable fraction of metal in the sediment. Such sequential extraction procedures suffer from potential shifts in the equilibrated distribution of metals between binding sites of different relative binding affinities during the extraction process, such that the final percentage distribution of metals measured is not necessarily that originally present in the sediment. What metal in a sediment is available trophically to a sediment-ingesting animal may well vary between deposit feeders according to interspecific differences in digestive physiology, and so it is difficult to define an absolute trophically available fraction. Nevertheless an easily extractable fraction (defined by the chemical nature of the extracting solution, e.g. a defined concentration of acetic acid or EDTA) may be a useful comparative model of what metal content of the sediment might potentially be bioavailable (to many deposit feeders), but care is needed not to refer to a specific chemically extracted fraction as the bioavailable fraction per se. 109

4 BIOMONITORS A biomonitor was defined above as an organism which accumulates trace metals in its tissues, the accumulated metal concentration of which may be analysed to provide a relative measure of the total amount of metal taken up by all routes by that organism, integrated over a preceding time period (Luoma and in press). Biomonitors, by definition, are net accumulators of trace metals ( 2002), typically to high body concentrations. These concentrations are easily measured with little significant risk of contamination during collection and laboratory processing. The accumulated metal derives from all routes of metal uptake, typically from solution (across permeable epithelial, typically respiratory, surfaces) and the diet (across the alimentary tract), in the case of aquatic invertebrates. Some of the metal taken up may be excreted, but nevertheless the accumulated metal concentration is an integrated measure of all the bioavailable supplies of metal to that organism - in short of the total bioavailability of that metal to that organism, over a preceding time period. A knowledge of the accumulation kinetics of a particular metal in a particular organism will define that time period; nevertheless, pragmatically, accumulated metal concentrations in specimens of the same species are typically used as comparative measures of total bioavailabilities of trace metals to that species without a precise time period being defined. The key point here is that the accumulated concentration of a metal in a biomonitor is an integrated measure of the metal taken up from all sources, i.e. an integrated measure of all bioavailable sources of metal to that biomonitor. This is absolutely the most direct measure of bioavailable metal to an organism, the very metal that is of ecotoxicological significance. It also follows that different organisms will reflect different routes of uptake to different degrees for the same metal. Strictly, then, a relative bioavailability needs to refer to a specific organism as well as to a specific metal. Again pragmatically, there are typically good correlations between high bioavailabilities of a metal to many biota resident in an aquatic habitat, and it is not uncommon to refer to a habitat as having a high bioavailability of a particular metal without specific reference to particular organisms. A more precise breakdown of relative metal bioavailabilities in different sources (e.g. solution, suspended material, sediment, etc.) can be made using a suite of different biomonitors, adding to the utility and value of the use of biomonitors to investigate toxic metal bioavailabilities in aquatic habitats. As well as being known to be (preferably strong) net accumulators of the metal concerned ( 2002), biomonitors should have several other desirable characteristics. Biomonitors should be sedentary to be representative of the study area, abundant, and easily identified and sampled. They should be large enough for analysis, even if necessary for the separation by dissection of a particular (metal storage) organ for analysis. A long-lived species has the potential to reflect bioavailabilities over a longer time period; so much the better if that species accumulation kinetics of a particular metal are known so that the time period reflected is identified. Biomonitors should be hardy, tolerant of raised metal bioavailabilities and of physicochemical variation in the habitat, for example changes in salinity in an estuary. They should also be tolerant of handling in the laboratory (not least to establish bioaccumulation kinetics) and in the field (for transplant experiments). ACCUMULATION PATTERNS Aquatic organisms differ in their patterns of accumulation of different trace metals ( 2002). At one extreme, all the metal taken up by an organism may be accumulated without significant excretion; at the other extreme, the rate of uptake of metal from all sources may be balanced by the rate of excretion so that there is no net accumulation over time. In the first example, the metal content (µg) of the organism is a direct measure of all metal taken up and therefore of the total bioavailability of the metal from all sources to that organism; the metal concentration (µg g -1 ) is similarly a measure of the metal taken up but will vary with weight changes, for example with growth (growth dilution), development or loss of gametes, build up or use of energy reserves. The process of the balancing of metal uptake and excretion rates is termed regulation ; this is common at the tissue and organ level in many invertebrates and vertebrates. Indeed many accumulators of trace metals will regulate trace metal concentrations in certain tissues (for example muscles) and store the majority of accumulated metal in one or more specific organs such as the hepatopancreas (digestive gland) or kidney. Such stored metal is invariably in detoxified form (bound to an insoluble (granule) or soluble (the protein metallothionein) binding site of high affinity), limiting release of the metal in a metabolically available form to exert a toxic action by binding in the wrong place ( 2002). Regulators of the metal content or concentration of the whole body regulate metal levels in all organs. Whole body regulation is restricted to essential, not non-essential, metals and only certain invertebrates, for example decapod crustaceans for zinc, appear to be able to carry out whole body regulation of trace metal concentrations ( 2002). The crustaceans are particularly good exemplars of the gradient of accumulation patterns found in aquatic invertebrates ( 1998, 2002, 2007). Barnacles are classic strong accumulators of many trace metals, for example zinc. Zinc taken up from solution by barnacles is accumulated without significant excretion ( and White 1989), and any excretion of zinc taken up from the diet is also extremely limited (1346 days half-life in Elminius modestus - and Wang 2001). Accumulated body concentrations therefore increase over the lifetime of the barnacle and can reach very high values (e.g µg Zn g -1 or more Table 2) in proportion to local Zn bioavailability, making barnacles very good biomonitors ( 1987, 1998; Phillips and 1988). The vast majority of the accumulated zinc is inevitably in detoxified form, in this case bound in zinc pyrophosphate granules (Walker et al. 1975a, b; Pullen and 1991). At the other end of the spectrum of accumulation patterns, the decapod crustacean Palaemon elegans regulates its body concentration of zinc (to about 90 µg Zn g -1 ) when exposed 110

5 Table 2. Variability in Zn concentrations among crustaceans. A selection of body concentrations of Zn in a systematic range of crustaceans from metal-contaminated (C) and uncontaminated (U) sites (from 2007; Luoma and in press). to a wide range of dissolved zinc bioavailabilities (White and 1982; and White 1989), an accumulation pattern for zinc also shown by other caridean decapods such as Pandalus montagui (ca 70 µg Zn g -1 ) (Nugegoda and 1988) (Table 2). The zinc uptake rate of P. elegans is significant (14% of total body Zn content per day at 100 µg Zn l -1 under specific physicochemical conditions at 20 C) and does change with Zn bioavailability, but the uptake rate is balanced by the excretion rate (White and 1982, 1984a, b). The body Zn concentration, therefore, remains unchanged over most environmentally realistic Zn exposures. Such decapod crustaceans cannot therefore be used as biomonitors of Zn bioavailabilities in the field. At a threshold dissolved Zn bioavailability, the excretion rate fails to match the uptake rate (regulation breakdown); there is then a net increase in body zinc concentration to only about double the regulated body concentration, but with lethal toxic effect (White and 1982). It appears, therefore, that much of the zinc in the body is in metabolically available form without detoxification. Thus, at high zinc uptake rates, the concentration of accumulated Zn in metabolically available form builds up sufficiently in the body to cause toxicity ( 1998, 2002, 2007). The accumulation pattern for zinc in amphipod crustaceans is intermediate, still one of net accumulation but much weaker accumulation than seen in barnacles ( and White 1989). Accumulated zinc is stored in detoxification granules in the ventral caeca of the alimentary tract ( 1998, 2002). These granules are excreted with the faeces at the end of the epithelial cell cycle in the caeca (Galay Burgos and 1998), and so the accumulated zinc concentration does not continue to increase as in barnacles. This is not a process of regulation for the body Zn concentration reaches a new steady state level as Zn bioavailabilities change, the bioavailability of Zn being reflected in the number of granules in (and hence the Zn concentration of) the ventral caeca. The amphipod body concentrations of Zn thus reflect ambient Zn bioavailability, and amphipod crustaceans are suitable trace metal biomonitors although their accumulated Zn body concentrations are much lower than those of barnacles (Table 2). Similar generalisations can be made about the accumulation of copper by crustaceans. Barnacles are strong net accumulators of copper, amphipods weaker net accumulators (with detoxification in granules in the ventral caeca), and caridean decapods appear to regulate body copper concentrations ( 1998, 2002, 2007). Decapods do not regulate body concentrations of any non-essential metal (e.g. cadmium) ( 2002), all crustaceans being net accumulators to different degrees with detoxification involving metallothioneins (Amiard et al. 2006). 111

6 BIOACCUMULATION KINETICS AND BIODYNAMIC MODELLING A knowledge of the bioaccumulation kinetics of a particular metal in a particular biomonitor increases the latter s utility by providing a time context over which metal uptake has been integrated. Furthermore, a knowledge of relatively few physiological parameters allows the application of biodynamic modelling which has the powerful advantage of allocating relative strengths to different metal uptake routes, for example solution versus diet (Luoma and 2005). The accumulated trace metal concentration of an aquatic invertebrate results from metal taken up from solution and metal taken up from food, after subtraction of any metal that has been excreted. If necessary, allowance can also be made for any growth of the animal over the time period investigated, for growth dilution will reduce the accumulated concentration. The rate of metal uptake from solution (µg g -1 d -1 ) can be calculated as k u C w, where C w is the dissolved metal concentration and k u is the dissolved metal uptake rate constant (µg g -1 d -1 per µg L -1, or L g -1 d -1 ). The rate of trace metal uptake from solution by an aquatic invertebrate is directly proportional to the dissolved concentration over the range of environmentally realistic dissolved concentrations, and the uptake rate constant k u is the slope of this relationship (Luoma and 2005). k u is measured in the laboratory for a defined set of physicochemical variables (e.g. salinity, temperature) of a medium, and it is then possible to calculate a metal uptake rate at any particular dissolved concentration. The rate of metal uptake from the diet can be expressed as AE (IR) C f where IR is the ingestion rate (g g -1 d -1 ), C f the concentration of the metal in the food (µg g -1 ), and AE the assimilation efficiency of the metal from that food source. AE can be measured relatively easily in the laboratory using radiotracers (Wang and Fisher 1999; and Wang 2001). The efflux rate constant k e (d -1 ) is also measured in the laboratory with radiotracers ( and Wang 2001). The growth rate constant g (d -1 ) allows for the effect of growth dilution. These separate terms can then be combined into the biodynamic model to predict the accumulated concentration of a trace metal in an aquatic invertebrate when exposed to a particular dissolved metal concentration and feeding on a diet with known metal concentration (Luoma and 2005). C = [k u C w + AE (IR) C f ] / [k e + g] The accumulation of zinc by barnacles can be used as an example of the application of the biodynamic model. Barnacles have a high rate of uptake of zinc from solution in comparison to other crustaceans ( and White 1989, 1998). Thus Eliminius modestus has a Zn uptake rate constant of 0.3 L g -1 d -1 at 10 C ( and White 1989, 1990), a Zn AE varying from 40% to 90% according to the food source and an efflux rate constant of d -1 ( and Wang 2001). Using data from the literature for Zn concentrations in water and food (suspended material), ingestion rates and growth rates, and Wang (2001) used the biodynamic model to predict an accumulated Zn concentration in E. modestus from Southend, England of 1500 to 4400 µg Zn g -1 according to the growth rate constant chosen. The measured concentration for barnacles collected in summer 2000 was 3463 ± 1155 µg Zn g -1 ( and Wang 2001), supporting the suitability of the biodynamic model as a model of trace metal bioaccumulation in the field (Luoma and 2005). Furthermore the model predicted that more than 97% of both accumulated Zn and Cd in E. modestus is accumulated from dietary ingestion ( and Wang 2001). Similarly Wang et al. (1999) showed using biodynamic modelling that both Zn and Cd are obtained for the most part from dietary ingestion rather than the dissolved phase in another barnacle, Balanus amphitrite, from Hong Kong coastal waters, again in spite of very high uptake rates from solution ( et al. 2003). Ultimately the high trace metal assimilation efficiencies and high ingestion rates of barnacles cause dietary uptake to predominate ( in press). The adoption of an accumulation pattern of storage detoxification with very limited excretion contributes to cause the high accumulated concentrations observed, with potential to increase further in circumstances of high metal bioavailability (Table 2). SUITES OF BIOMONITORS Reference has been made to different uptake routes for trace metals. For aquatic invertebrates these reduce to uptake from solution (typically the water column) and uptake from the diet. The biology (e.g. diet, epifaunal or infaunal habit) of the organism chosen as a biomonitor will determine the sources of metal available to that biomonitor (relative contributions of sources being assessable by biodynamic modelling). This is a strength of the use of a suite of biomonitors in a habitat, the choice of biomonitors allowing the sampling of different trace metal sources in a habitat. Potential suites of biomonitors of trace metal bioavailabilities in coastal waters are provided in Table 3 for both European and (more hypothetically) Australian coastal waters. Macrophytic algae lack the roots of angiosperms like sea grasses and take up metals only from solution (so long as fronds are not in direct contact with sediment). Such algae are therefore excellent biomonitors of the bioavailable metal in solution in the water column (with the caveat that uptake by algae is not necessarily absolutely identical to that by invertebrates, which themselves may show subtle variation in the relative importance of different routes of trace metal uptake from solution (Luoma and in press)). Care must be taken to sample the frond at a pre-determined distance from the distal end (10 cm for Fucus vesiculosus) to allow time for new growth to equilibrate with the environment, and the frond surface should be free of epiphytes (Bryan et al. 1985; et al. 2002). F. vesiculosus is a widespread littoral brown alga in Europe with a literature on accumulated metal concentrations in different habitats (Bryan and Gibbs 1983; et al. 2002). An ecological analogue in Australia may be a species of the brown alga Sargassum. The two green algal genera Ulva and Enteromorpha are widespread; species of both are used as biomonitors although specific identifications are not easy and therefore may not be always reliable ( and Phillips 1993; 1995b). 112

7 Table 3. Possible suites of trace metal biomonitors suitable for European and Australian coastal waters. Mussels and oysters are commonly used for trace metal biomonitoring in coastal waters ( and Phillips 1993; Phillips and 1994; 1995b). These epifaunal bivalves obtain their metals from solution in the water column and from the suspended food particles that they filter with their gills (Table 3), the two sources inevitably being correlated to some degree. Different bivalve species potentially filter feed on slightly different ranges of particle size, providing subtle differences in the food source and therefore the potential metal sources sampled. Mussels and oysters usually filter suspended detritus particles and phytoplankton, near the bottom end of the potential size range of plankton, the smaller sized particles being more likely to be organically- and metal-rich than their zooplankton counterparts. Barnacles on the other hand will filter out zooplankton. There are again subtle differences between species as to the size range filtered. All species of barnacles feed by captorial feeding on larger particles such as zooplankton using the three posterior pairs of thoracic limbs (cirri); species of Balanus (and Tetraclita squamosa) also feed by the filtration of small suspended particles (microfeeding) in the size range of detritus and phytoplankton using the setae of the anterior cirri (Crisp and Southward 1961; Anderson 1981, 1994; Hunt and Alexander 1991). It is probable that individual barnacle species feed on slightly different size ranges of suspended particles, or at least to different extents on different parts of the size spectrum available (Anderson 1981, 1994). Suspension-feeding lamellibranch bivalves and barnacles therefore sample the water column and suspended particles as metal sources (Table 3). In European waters the mussel Mytilus edulis is well established as a trace metal biomonitor with a good literature on comparative accumulated trace metal concentrations (Luoma and in press). Related mussel species M. trossulus and M. galloprovincialis are also used 113

8 in biomonitoring programmes ( and Phillips 1993; 1995b; Szefer et al. 2006), and M. galloprovincialis has been introduced to the cooler waters of Australia (Seed 1992) (Table 3). The European oyster Ostrea edulis can be used but the Pacific oyster Crassostrea gigas, spread worldwide by man for mariculture (including to Australia), is more commonly used for biomonitoring (Table 3). The indigenous Australian rock oyster Saccostrea cucullata (also referred to as S. commercialis and S. glomerata) is also a suitable biomonitor, as is its congener S. echinata in warmer Australian waters (Table 3). The blood cockle Anadara trapezia burrows in intertidal mud in sheltered habitats and has considerable potential as an Australian trace metal biomonitor. Amongst the barnacles (Table 3), Balanus improvisus lives in low salinity coastal habitats in Europe, and has been used, for example, as a biomonitor of contaminant trace metals in the Thames estuary ( et al. 2002) and in the Gulf of Gdansk in the Baltic receiving effluent from the Vistula river ( et al. 2004). The fouling barnacle Balanus amphitrite is cosmopolitan in the tropics and subtropics, and can be collected in both Europe and Australia. Elminius modestus has been moved from south to north anthropogenically and is another potential biomonitor in both hemispheres (Table 3). Sediments act as sinks for trace metals in estuaries and other coastal waters and are a source of metals for animals, particular burrowing infauna. It is important to distinguish two routes of metal uptake from the sediment firstly dietary uptake of metals associated with sediment ingested by deposit feeders, and secondly uptake from solution in interstitial pore water released in equilibrium from the sediment particles. For soft-bodied invertebrates (e.g. polychaete worms) burrowing in sediments, the strength of any irrigation current will determine the relative contributions of overlying water column water and interstitial pore water to the water bathing the animal in the burrow and serving as the source of dissolved metal. For infaunal lamellibranch bivalves, the shell and mantle edges will provide a barrier to isolate the animal s gills from the interstitial water, and the dissolved source of metal will typically be the overlying water column via the irrigation current. Lucinid bivalves with symbiotic chemosynthetic bacteria in the gills will differ from this scenario and are discussed further below. It is clearly apparent that a knowledge of the biology of a potential biomonitor is crucial to understanding the potential routes of metal uptake available to that biomonitor. Tellinid bivalve molluscs typically burrow in the sediment and ingest surface deposited particles by sucking them up with the inhalant siphon. The water entering the inhalant siphon is derived from the water column and also serves as a respiratory current delivering oxygenated water to the gills. Thus tellinids potentially sample the water column and surficial sediments for metals. The pioneering work of Geoff Bryan and colleagues in the metal-rich estuaries of southwest England highlighted the suitability of two tellinid bivalves, Macoma balthica and Scrobicularia plana (Table 3) as biomonitors of the bioavailabilities of trace metals in estuarine sediments (Luoma and Bryan 1978, 1982, Bryan et al. 1980, 1985, Bryan and Gibbs 1983). In Australian waters Tellina deltoidalis may be a suitable equivalent (Table 3). Burrowing polychaetes (or indeed other worms such as oligochaetes, sipunculids, etc. or holothuroid echinoderms) are soft bodied and therefore have the potential to take up metals from the water bathing them in the burrow. Whether this water consists only of interstitial pore water or is affected to any degree by irrigation water drawn down from the water column by the animal will vary with species. Terebellid polychaetes typically have their bodies bathed in burrow water while they feed on surficial sediment material with radiating tentacles. Sabellid polychaetes, on the other hand, may withdraw down tubes into a burrow but also sample water column water and filter feed on suspended particles with a crown of tentacles. Table 3 suggests some potential candidates as biomonitors to sample these different sources. Burrowing invertebrates that ingest deeper sediments include polychaetes (Table 3). A lot of work has been carried out on the estuarine ragworm Nereis diversicolor as a trace metal biomonitor of sediments particularly again by Geoff Bryan in SW England (Luoma and Bryan 1978, 1982, Bryan et al. 1980, 1985, Bryan and Gibbs 1983). The lugworm Arenicola marina processes much sediment through its gut, and, with gills down the body, likely takes up metal from bathing water that is influenced by both irrigation current and sediment interstitial water. Lucinid bivalves offer a unique opportunity to sample dissolved metal sources without any interference from dietary metal uptake. Lucinids have symbiotic chemoautotrophic bacteria in the gills. The bivalve supplies these bacteria with a current of interstitial sulphide-rich water as an energy supply for chemosynthesis and the bacteria supply organic compounds as a food source for the bivalve (Distel 1998). The bivalve also needs a separate current of oxygenated water column water for its own respiratory needs. These two supplies of water will also provide bioavailable metals in solution, the bioavailable supplies expectedly differing greatly between the reduced and oxygenated water supplies with their different metal concentrations and physicochemical characteristics. In a pioneering study, Silva et al. (2006) used the lucinid Phacoides pectinata as part of a suite of biomonitors in mangrove-lined estuaries of Brazil. Potential lucinid bivalve metal biomonitors are Loripes lucinalis in Europe, and Wallucina assimilis (Glover and Taylor 2001) and Austriella corrugata in Australia (southern and northern coasts respectively) (Taylor and Glover personal communication). Talitrid amphipod crustaceans are often easily collected from strand lines and are of a convenient size for individual analysis. Talitrids take up metals from both solution and food (Weeks and 1991, 1993) which consists of the cast up plant matter typically macrophytic algae. Such algae receive their metals from solution, and it is arguable that talitrid crustaceans are biomonitoring predominantly the bioavailable metal in the water column. The common European talitrid Orchestia gammarellus has been used as 114

9 Table 4. Potential cosmopolitan trace metal biomonitors and their geographical distributions, often spread anthropogenically (man). a trace metal biomonitor around the UK ( et al. 1989, Moore et al. 1991) including the Thames estuary ( et al. 2002). Another talitrid, the sand hopper Talitrus saltator, proved to be the only reasonably accessible invertebrate biomonitor on the tideless shores of the Baltic (Fialkowski et al. 2000). Talitrids (particularly Transorchestia chiliensis) have also been used as trace metal biomonitors in New Zealand (Marsden et al. 2003), and a potential talitrid amphipod biomonitor in Australia is Talorchestia cavimana (Table 3). COSMOPOLITAN BIOMONITORS Reference has been made above to the availability of literature databases of accumulated trace metal concentrations in particular species in order that a measured concentration can be interpreted as high or low in context. The more widespread the distribution of a single biomonitoring species, the greater its value as a cosmopolitan biomonitor providing cross-reference across large geographical areas ( and Phillips 1993, Luoma and in press). Table 4 suggests several potential candidates as cosmopolitan trace metal biomonitors in coastal waters, using the categorisation of metal sources employed in Table 3. Many gaps still exist and, ironically, many of the best candidates are species spread anthropogenically, either as fouling species (e.g. the barnacle Balanus amphitrite) or for mariculture (the mussels Mytilus edulis and M. galloprovincialis, the oyster Crassostrea gigas and the clam Ruditapes philippinarum). The fouling barnacle B. amphitrite has been spread by shipping worldwide in tropical and subtropical waters. It has the advantage of growing on manmade installations (piers, etc.) in still waters typical of ports receiving industrial effluent the very areas that might suffer metal pollution. Based on work in Hong Kong, there is a good reference database of accumulated metal concentrations in this barnacle (Table 5) ( and Blackmore 2001). The worldwide distribution of B. amphitrite means that it occurred at the mouth of an estuary in Natal, Brazil, and the Hong Kong reference database allowed Silva et al. (2006) to benchmark the ambient bioavailability of trace metals to barnacles in that Brazilian estuary as low (Table 6). Mussels, particularly species of Mytilus, are the classic trace metal biomonitors (Phillips and 1994, 1995b) having long been used in Mussel Watch programmes in the USA (Lauenstein et al. 1990) and also elsewhere in the world (Cantillo 1998, Szefer et al. 2006). The two commonly 115

10 Table 5. Weight-adjusted mean accumulated concentrations of trace metals (µg g -1 dry weight) in the bodies of barnacles, Balanus amphitrite, from Hong Kong coastal waters in April 1998 (from and Blackmore 2001). Concentrations in bold can be considered high and indicative of atypically raised bioavailability of that metal in the local habitat. Table 6. Balanus amphitrite: Comparative metal concentrations in 4.5 mg barnacle bodies as estimated from best-fit double log regressions log M = log a + b log W) in barnacles from the Curimataú estuary, Natal, Brazil (with 95% Confidence Limits) (Silva et al. 2006) and in barnacles from 12 Hong Kong sites (range) ( and Blackmore 2001). (After Silva et al. 2006) Table 7. Mean accumulated concentrations of trace metals (µg g -1 dry weight) in the soft tissues of the mussel, Perna viridis, in Hong Kong coastal waters in April 1986 (from Phillips and 1988). Concentrations in bold can be considered high and indicative of atypically raised bioavailability of that metal in the local habitat. used mussels, M. edulis and M. galloprovincialis, have been spread round the world, although they are not easy to tell apart from each other, nor from their northern counterpart M. trossulus (Seed 1992, and Phillips 1993). Less widespread but still of great value because it occurs throughout the IndoPacific is the green mussel Perna viridis, while its close relative Perna perna is found in Africa and along the Atlantic coast of South America (Table 4) (Siddall 1980, and Phillips 1993, 1995b). Part of the reference database for P. viridis (based on Hong Kong data Phillips and 1988) is shown in Table 7, providing a context for measured metal concentrations in this mussel. While rock oysters of the genus Saccostrea are notoriously difficult to identify, another oyster, Crassostrea gigas, is large, conspicuous and rarely misidentified. It is now used in mariculture worldwide and therefore has great potential as a cosmopolitan biomonitor. Oysters also have the advantage over mussels in showing stronger net accumulation of many trace metals than mussels ( 2002, Geffard et al. 2004). The clam Ruditapes philippinarum is more widespread than its specific name suggests; data are building up, on field metal concentrations and on biodynamic modelling of their accumulation (Chong and Wang 2001). A widespread lucinid bivalve in the IndoPacific is Codakia tigerina, while C. orbicularis is a Caribbean (Atlantic) equivalent (Taylor and Glover personal communication). The talitrid amphipod Platorchestia platensis has been reported worldwide and can be used as a trace metal biomonitor ( and Phillips 1993, 1995b). 116

11 SYSTEMATIC IDENTIFICATION Invertebrates show a wide range of trace metal accumulation patterns and correspondingly different invertebrates collected from the same habitat show very different accumulated metal concentrations. Differences between accumulation patterns for different trace metals in just the crustaceans have been emphasised above and detailed in Table 2. It is therefore meaningless to compare metal concentrations between organisms of different taxa. As can be seen from mussels (Perna viridis) and barnacles (Balanus amphitrite) collected from the same waters (Tables 5 and 7), what can be considered a high concentration in the two species can differ by one or two orders of magnitude. A high Zn concentration in the barnacle is µg g -1, in the mussel 168 µg g -1 ; a high Cu concentration is 1810 µg g -1 in the barnacle, 219 µg g -1 in the mussel (Tables 5 and 7). The question then arises how closely related must two biomonitors be before their trace metal accumulated concentrations can be compared? The bottom line is that, unless specifically shown otherwise, it is probably invalid to make comparisons even between species of the same genus. Lobel et al. (1990) collected two mussels, Mytilus edulis and M. trossulus, from the same uncontaminated site at Bellevue, Newfoundland, Canada. While there were no significant differences in accumulated concentrations of As, Cu, Mn, Pb, Se, V and Zn, there were for Ag, Cd, Mo and U (Lobel et al. 1990). Similarly there were significant differences in the accumulated concentrations of Cu in the barnacles Balanus amphitrite and Balanus uliginosus collected simultaneously from the same sites near Xiamen, China (Table 8) ( et al. 1993b). Related talitroidean amphipods including the congeneric species Orchestia gammarellus and O. mediterranea, all collected from the Isle of Cumbrae in the Firth of Clyde, Scotland, also show significant interspecific differences in Cu and Zn concentrations (Moore and 1987). Thus it is crucial to be sure of specific identifications of biomonitors unless it has been specifically shown that potentially confused species do not show significant interspecies differences in accumulated metal concentrations. This is the case for example for two sympatric freshwater mayfly larvae, Baetis rhodani and B. vernus in metalcontaminated streams of Upper Silesia, Poland (Fialkowski et al. 2003). EXAMPLES OF THE USE OF TRACE METAL BIOMONITORS The first use of a suite of biomonitors to assess ambient trace metal bioavailabilities in a coastal habitat can be attributed to Geoff Bryan and his research collaborators in the 1970s and 1980s (Bryan et al. 1980, 1985). The original suite of biomonitors used in the estuaries of SW England, affected by metal-rich sediments derived from massive mining activity in the 19 th Century, consisted of the tellinid bivalves Scrobicularia plana and Macoma balthica, and the polychaete Nereis diversicolor (Bryan et al. 1980). A comprehensive study of the potential of many estuarine organisms as metal biomonitors (Bryan and Gibbs 1983; Bryan et al. 1985) concluded that a reasonable monitoring programme should Table 8. Weight-adjusted mean Cu concentrations (µg g -1 dry weight, with 95% Confidence Limits) in the bodies of barnacles Balanus amphitrite and B. uliginosus as estimated from best-fit double log regressions, collected in 1991 in the region of Xiamen, Fujian Province, China (after et al. 1993b). involve the analysis of several species including perhaps a seaweed (e.g. Fucus vesiculosus), a suspension feeder (e.g. Mytilus edulis), a deposit feeder (e.g. Scrobicularia plana) and a carnivore (e.g. Platichthys flesus), (Bryan et al. 1985). This precocious study investigated a series of correlations between metal concentrations in different biomonitors and in sediments, as it defined what metal source was being monitored by what biomonitor (Bryan et al. 1985). Bryan and Langston (1992) later used these comprehensive studies as the basis of a review of the biomonitoring of metals associated with sediments, particularly in estuaries. In the Far East, Phillips and (1988) combined their expertise to compare the green mussel Perna viridis and three species of barnacles, Balanus amphitrite, Tetraclita squamosa and Capitulum mitella as biomonitors of trace metals in Hong Kong waters. Hong Kong proved to be an excellent location to test the different capacities of organisms to act as trace metal biomonitors because its waters had a well defined gradient of metal contamination, and P. viridis, B. amphitrite and T. squamosa are widespread in the IndoPacific (and beyond in the case of B. amphitrite). Although each species accumulated different absolute concentrations of metals, agreement between rank orders of sites along contamination profiles exhibited by the four species was excellent for chromium, copper, lead and zinc (but not, incidentally, for cadmium, see below) (Phillips and 1988). Thus the four suspension feeders were presumably identifying variation in the same bioavailable sources of the four trace metals, dissolved and via suspended detritus and plankton, even though P. viridis might be expected to take more of the smaller (more organically- and metal-rich) suspended particles than the barnacles (Phillips and 1988), and B. amphitrite would perhaps take more smaller particles than in turn T. squamosa and C. mitella (Anderson 1981, 1994). This study carried out in 1986 (Phillips and 1988) investigated the spatial variation of comparative metal bioavailabilities across Hong Kong waters, but also provided a baseline against which future biomonitoring programmes could follow changes in local metal pollution over time. After a second survey in 1989, a comprehensive survey was carried out at the same time of year in 1998 involving both B. amphitrite and T. squamosa to analyse geographical variation 117