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Biological Activity of Metals FINAL REPORT Caroline Whalley, Juan Brown, Mark Kirby, Liam Fernand DETR/DEFRA contract CDEP/84/5/285 June 2002

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Biological Activity of Metals

FINAL REPORT

Caroline Whalley, Juan Brown, Mark Kirby, Liam Fernand

DETR/DEFRA contract CDEP/84/5/285

June 2002

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Metal Biological Activity Final Report

Executive Summary

The purpose of this project was to investigate associations between cadmium and phytoplankton in marine sediments in the central North Sea, and to consider the OSPAR JAMP (Joint Assessment and Monitoring Programme) recommendations for the application of biological effects methods for metals in marine waters.

The work was required to investigate:

concerns regarding the cause of elevated concentrations of cadmium and chromium in the sandy sediments of the Dogger Bank.

the applicability in offshore UK marine waters of the JAMP recommended cascade of biological effects techniques for metals.

An example of why it was important to carry out this work is provided by international concerns about the quality of the environment around the Dogger Bank. The QSR (1993) identified the possibility that the Dogger Bank is affected by transport of nutrients and contaminants from coastal regions. To support decisions about possible action in the case of environmental harm, we need to know more about the causes and processes that affect observations. Recent MAFF-funded work had described a summertime circulation connecting the north-east coast of England and the northern flanks of the Dogger Bank. Within these waters it was shown that there was an association between particulate cadmium and chlorophyll a concentration. Consequently, it was possible that cadmium concentrations within sediment were linked to phytoplankton production rather than any advection from the coast. Bioavailability of cadmium in Dogger sediments is of interest since dab (Limanda limanda) livers collected from the Bank have shown relatively high cadmium concentrations in the past (CEFAS, 1998).

The principal findings of the project were

Of several benthic species considered, Echinocardium cordatum showed most promise for metallothionein induction of in-situ species, although further work is required to better understand MT dynamics in this organism.

To test whether phytoplankton deposition affected cadmium concentrations in the sediment, surface material from sediment cores collected at the same sites in summer 2000 and winter 2001 in the Dogger region was analysed. As all cadmium concentrations were below the detection limit (0.05 mg kg–1) identification of a link between chlorophyll a and cadmium in sediments was not possible.

Sediment cores from a transect across the bank during 1999 showed elevated concentrations of cadmium (up to 0.34 mg kg-1 ) and chromium (up to 78 mg kg-1) from 5 cm and 10 cm depths on top of the Bank, but not in the surface (top 1 cm) sediment layer. Concentrations were not elevated in the finer sediments either side of the Bank.

Evidence for direct transport of contaminants via the seasonal jet-like circulation from coastal waters was not found. Elevated concentrations of cadmium in deeper sediments were present across the whole of the Bank.

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Contents

Executive summary

Table of contents 3 List of figures 5

1) Introduction and background to the work1.1: Introduction 7 1.2 : Context of work on the Dogger Bank 7 1.3 : References 13

2) The JAMP Cascade of Biological Effects Methods for Metals - Literature review

2.1: Introduction 14 2.2: Lysosomal Stability 14

2.2.1 Introduction 14 2.2.2 Analytical Methods 15 2.2.3 Field Applications and Practicalities 16 2.2.4 Influences on the Biomarker 17 2.2.5 Discussion - Lysosomal stability 18

2.3: Oxidative Stress 19 2.3.1 Introduction 19 2.3.2 Analytical Methods 19 2.3.3 Field Applications and Practicalities 19 2.3.4 Influences on the Biomarker 20 2.3.5 Discussion - Oxidative stress 20

2.4: Metallothionein 21 2.4.1 Introduction 21 2.4.2 Analytical Methods 21 2.4.3 Field Applications and Practicalities 23 2.4.4 Influences on the Biomarker 24 2.4.5 Discussion – Metallothionein 26

2.5: -Amino Levulinic Acid Dehydratase (ALA-D) 27 2.5.1 Introduction 27 2.5.2 Analytical Methods 27 2.5.3 Field Applications and Practicalities 28 2.5.4 Influences on the Biomarker 29 2.5.5 Discussion - ALA-D 30

2.6: Conclusions and Recommendations 31 2.7: References 32

3) Investigation into metallothionein induction in benthicorganisms collected between Tyne/Tees and western Dogger Bank.

3.1 : Introduction 38 3.2 : Methods 38 3.3 : Analytical quality control 40 3.4 : Results 40

3.4.1 Metals in sediments 40

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3.4.2 Metallothionein results 41 3.5 : Discussion 43 3.6 : Conclusions 44 3.7 : References 44

4) Marine phytoplankton – a mechanism for transferring dissolved cadmium to sediments – Literature review

4.1 : Summary 45 4.2 : Introduction 45 4.3: Nutrients in seawater 46 4.4 : Plankton dynamics 46 4.5 : Toxicity of cadmium to phytoplankton 48 4.6 : Cadmium as a nutrient 48 4.7 : Cadmium uptake by phytoplankton 49 4.8 : Trophic transfer and sedimentation 51 4.9 : Metal flux by plankton blooms 53 4.10 : Production in the Dogger Bank region 53 4.11 : Cadmium in sediments from the Dogger Bank region 55 4.12 : Cadmium in benthos from the Dogger Bank region 55 4.13: Conclusions 55 4.14 : References 56

5) Investigations into cadmium and chlorophyll a concentrations in sediments from the Dogger Bank region

5.1 : Introduction 60 5.2 : Practical work 60 5.3 : Analytical Quality Control 63 5.4 : Results 63

5.4.1 : Seasonal sediment metal concentrations 63 5.4.2 : Discussion of results from seasonal sediment 71

sampling5.4.3 : Dissolved cadmium in waters off the north east 74

coast and from the Dogger dogleg transect5.4.4 : Metals in cored sediments from the Dogger Bank 74

dogleg, 19995.4.5 : Considering possible pathways for cadmium to 84

reach sediments at the Bank.5.6 : Discussion 85 5.7 : Further work 86 5.8 : References 87

Annex 1 : Preparation of biological samples 88 Annex 2 : Analytical procedures for metallothionein and protein 89

analysesAnnex 3 : Collection, preparation and analysis of sediment samples 92 Annex 4 : Recoveries from certified reference materials (CRMs) 93 Annex 5 : Sediment metal and chlorophyll concentrations, Dogger 94

Bank 2000-01

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List of figures

Fig 1.1 Outline of project structure 8 Fig 1.2 Cadmium concentrations in sediments (JMG survey) 9 Fig 1.3 Cadmium concentrations in sediments (Dogger Bank survey, 11

1994-95)Fig 1.4a Cadmium vs chlorophyll a in suspended particulates from the 12

Western North Sea and Dogger Bank, summer 1998Fig 1.4b Cadmium vs chlorophyll a in suspended particulates from the 12

Dogger Bank (DB) and Tyne Tees regions, summer 1999Fig 3.1 Sampling locations 39 Fig 3.2a MT induction in Asterias rubens 42 Fig 3.2b MT induction in Echinocardium cordatum 42 Fig 4.1 Diagrammatic representation of phytoplankton succession 47

sequence and associated physical-chemical changes (Lewis, 1979)

Fig 5.1a Scanfish section August 2000 61 Fig 5.1b Scanfish section Dogger dogleg (1999) 62 Fig 5.2 Comparison of recovery from sediment CRMs (BCSS-1 and 64

MESS-2)Fig 5.3a Al in surface sediment, Dogger transect 2000-01 66 Fig 5.3b Fe in surface sediment, Dogger transect 2000-01 66 Fig 5.3c Cr in surface sediment, Dogger transect 2000-01 67 Fig 5.3d Mn in surface sediment, Dogger transect 2000-01 67 Fig 5.3e Pb in surface sediment, Dogger transect 2000-01 68 Fig 5.3f Zn in surface sediment, Dogger transect 2000-01 68 Fig 5.4a Cr, Pb, Zn vs aluminium in sediments, Dogger transect 69

2000-01Fig 5.4b Mn vs aluminium in sediments, Dogger transect 2000-01 69 Fig 5.4c Fe, Li vs aluminium in sediments, Dogger transect 2000-01 69 Fig 5.5a Metals (Cr, Pb, Zn) vs iron in sediments, Dogger transect 70

2000-01Fig 5.5b Manganese vs iron in sediments, Dogger transect 2000-01 70 Fig 5.5c Metals (Al, Li) vs iron in sediments, Dogger transect 2000-01 70 Fig 5.6a Chlorophyll a in Dogger transect surface sediments – 72

summer 2000Fig 5.6b Chlorophyll a in Dogger transect surface sediments – 72

winter 2001Fig 5.7a Total pigment in Dogger transect surface sediments – 73

summer 2000Fig 5.7b Total pigment in Dogger transect surface sediments – 73

winter 2001Fig 5.8a Dissolved Cd along Dogger dogleg transect 1999 75 Fig 5.8b Dissolved Cd along Tyne Tees transect 1999 75 Fig 5.9a Dissolved Cd vs chlorophyll a in Dogger dogleg transect 76

1999Fig 5.9b Dissolved Cd vs chlorophyll a in Tyne/Tees transect 1999 76 Fig 5.10a Dissolved Cd vs Cd in suspended particulates in Dogger 77

dogleg transect 1999Fig 5.10b Dissolved Cd vs Cd in suspended particulates in Tyne/Tees 77

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transect 1999Fig 5.11a Dogger dogleg 1999 core data – Aluminium 79 Fig 5.11b Dogger dogleg 1999 core data – Iron 79 Fig 5.11c Dogger dogleg 1999 core data – Lithium 80 Fig 5.11d Dogger dogleg 1999 core data – Manganese 80 Fig 5.11e Dogger dogleg 1999 core data – Cadmium 81 Fig 5.11f Dogger dogleg 1999 core data – Zinc 81 Fig 5.11g Dogger dogleg 1999 core data – Chromium 82 Fig 5.11h Dogger dogleg 1999 core data – Lead 82 Fig 5.12a Cr vs Al in cored sediment Dogger dogleg 1999 83 Fig 5.12b Pb vs Al in cored sediment Dogger dogleg 1999 83 Fig 5.12c Cr vs Mn in cored sediment Dogger dogleg 1999 83 Fig 5.12d Cd vs Fe in cored sediment Dogger dogleg 1999 83

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1) Introduction and Background to the work

1.1 : Introduction

This project has two main themes (fig 1.1). One was to investigate the series of biological effects methods for metals, suggested by the OSPAR Commission as part of JAMP in 1998 (OSPAR, 1998), for their effectiveness in examining metal contamination effects in biota in marine waters. The other was to further investigate the anomalous cadmium concentrations found in some sediments from the Dogger Bank, following previous MAFF funded work (AE1214) which suggested a possible link between phytoplankton and cadmium concentrations within the water column. A further impetus to examine the Dogger Bank anomaly arose following the discovery of a seasonal (May – October) jet-like circulation from the north east English coast and skirting the northern side of the Dogger Bank (Brown et al, 1999, 2001).

The main aims of the project were:

To perform a literature review focused on the practical application of the series of linked analytical and biological effects techniques for metals suggested by OSPAR.

Using those biological effects techniques identified in the literature review as being robust and cost effective, to assess the biological activity of metals at a selection of offshore sites.

To investigate anomalous cadmium concentrations in sediments from the Dogger Bank, and test whether increased concentrations of contaminants there may result from natural, in situ processes.

To explore whether contaminant concentrations in sediment vary with the seasonal cycle of phytoplankton production and if the spatial variability is related to the distribution of primary production.

The two parts of the project proceeded in parallel, and are reported in different chapters in this report. Literature reviews were produced for both strands and are presented in chapters 2 and 4.

This work has benefited by being associated with other MAFF/DEFRA-funded projects, notably AE1214, AE1219, AE1225 and BEQUALM.

1.2 : Context of work on the Dogger Bank

Interest in metals in sediments at the Dogger Bank – particularly cadmium and chromium – was heightened in the early 1990s when the JMG survey (Rowlatt and Lovell, 1994a and b) found indications of some elevated concentrations in the sandy sediments. Such concentrations were unexpected because sands do not normally accumulate contaminants. Figure 1.2 shows cadmium concentrations in sediments collected from the region during the JMG survey. Concern was raised given the existence of industry on the North east English coast which discharged cadmium and chromium, although at the time there was no evidence for a pathway for the

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Fig 1.1

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fig 1.2(mapinfo plot not available in e-version)

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contaminants out to the Bank. Dab collected under the NMP from the western Dogger and the Tyne/Tees regions showed relatively high concentrations of cadmium in liver in samples collected during 1994 (CEFAS, 1998).

The internationally-based ICES/IOC Bremerhaven workshop examined a transect from the German Bight to the south eastern Dogger Bank. Results in cadmium concentrations in sediments and biota were varied (Cofino et al 1992). Sediments (<63µm) showed decreasing Cd concentrations with increasing distance from the coast. However, Cd concentrations in Aphrodite aculeata showed maximum values offshore, and those in dab liver and Pagurus bernhardus increased with increasing distance from the coast. Highest metallothionein (MT) concentrations were found in dab collected near the Dogger Bank, in contrast to that expected which was closer to the coast where sediment Cd concentrations were highest (Hylland et al, 1992). The elevated Cd and MT concentrations at offshore stations were thought to imply metal inputs in addition to those from the Weser and Elbe estuaries considered in the project.

Following the JMG survey, a project jointly funded by MAFF and DoE examined conditions at the Dogger in greater depth. This included a detailed sediments and benthos survey across the Bank region (Whalley et al 1997). This study confirmed that there were some elevated cadmium concentrations in the sediments (fig 1.3), but found less evidence for elevated chromium concentrations. It was noticeable that the “hot” spots did not occur in exactly the same locations as had been found during the JMG survey, prompting the comment that they “were consistently variable” (Rowlatt, pers. comm.). Examination of replicate grab samples, performed by D Lovell during the JMG survey, showed that while reproducibility of the entire sediment sampling and analysis procedure in samples was generally good, those for Cd in samples from the Dogger were considerably more variable. For example, Cd in 6 replicate grabs from the Dogger Bank (#775) showed a mean concentration of 74 µg kg -1, 95% confidence interval (CI) of 80%, while for German Bight samples (#781), the concentration was 36 µg kg-1 with 95% CI 30%. Reproducibility was much better for other metals, typically being 5-20%.

More recently, Brown et al (1999, 2001) discovered that seasonal stratification in the deeper waters (~40m) off the north east coast leads to a transport pathway during the spring-summer months. The pathway runs from the coast and skirts past the northern flank of the Dogger Bank. As part of this MAFF-funded investigation (AE1225), we examined the possibility that the pathway might transport particulate contaminants offshore. To this end, we collected 100l water samples at three depths from the water column (~6m below the surface, at the thermocline 20-30m depth, and ~6m above the seabed) using a CTD. The samples were filtered, retaining the particulate material for total digestion (HF) and metals analysis. Other analyses upon samples collected on the same CTD “dip” included chlorophyll a, which is used as a proxy for the phytoplankton content. Particulate samples were analysed for total metals content, which allowed normalisation for the clay content.

While normalisation for clays generally works well for most metals, it does not usually produce strong associations with cadmium. However, an unexpected relationship was found between the cadmium in the suspended load and chlorophyll a (fig 1.4). A cursory examination of the literature revealed evidence to suggest that

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Fig 1.3(mapinfo plot not available in e-version)

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Fig 1.4

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such a relationship might indeed occur, and if it did, this could have significant consequences upon our understanding of cadmium cycling in marine waters. At this stage, MLIS - DETR funded the current project.

1.3 : References Brown J, Hill AE, Fernand L and Horsburgh KJ. (1999) “Observations of a Seasonal Jet-like Circulation at the Central North Sea Cold Pool Margin” Est Coast Shelf Sci 48 (3) 343-355

Brown J, Fernand L, Horsburgh KJ, Hill AE. and Read JW. (2001). “Paralytic shellfish poisoning on the east coast of the UK in relation to seasonal density-driven circulation” J Plankton Res 23 105-116

CEFAS. (1998) Monitoring and surveillance of non-radioactive contaminants in the aquatic environment 1995 and 1996. CEFAS, Lowestoft. AEMR 51 ISSN 0142 2499

Cofino, WP, Smedes F, de Jong SA, Abarnou, A, Boon JP, Oostingh I, Davies IM, Klungsoyr J, Wilhelmsen S, Law RJ, Whinnett JA, Schmidt D and Wilson S. (1992) "The chemistry programme" Mar Ecol Prog Ser 91 (1-3) 47-56.

Hylland K, Haux C and Hogstrand C (1992) “Hepatic metallothionein and heavy metals in dab Limanda limanda from the German Bight” " Mar Ecol Prog Ser 91 (1-3) 89-96

OSPAR (1998) JAMP guidelines for contaminant-specific biological effects monitoring. OSPAR, London 38pp

OSPAR (1993) North Sea Quality Status Report OSPAR, London 132pp

Rowlatt S and Lovell D. (1994a) Survey of contaminants in coastal sediments. MAFF, Burnham-on-Crouch. DoE research contract PECD 7/7/358

Rowlatt S and Lovell D. (1994b) “Lead, zinc and chromium in sediments around England and Wales” Mar Poll Bull. 28 (5) 324-329

Whalley C, Rowlatt S, Jones L, Bennett M and Campbell S. (1997) Metals in sediments and benthos from the Dogger Bank, North Sea. CEFAS Burnham-on-Crouch, DoE contract CW0 301

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2) The JAMP Cascade of Biological Effects Methods for Metals- A Literature Review

2.1 : Introduction

In recent years there has been a trend towards associating biological effects measurements with chemical measurements of relevant media, such as the water or sediment in which an organism lives. This development has arisen since chemical or biological studies could be indicating a problem while being unable to identify but a causal association. For instance, biological effects data might suggest that something was causing deleterious changes in marine species, but could not be used to identify the particular cause of the problem. Meanwhile, chemical data might show the presence of contaminants but could not be used to identify any biological significance. The need to better understand the links between contaminant presence and biological effects has been highlighted in various fora, such as at OSPAR/ICES workshops.

In February 1998 the OSPAR commissions, as part of their Joint Assessment and Monitoring Programme (JAMP), issued guidelines for general biological effects monitoring and also for contaminant-specific biological effects monitoring (OSPAR, 1998). Within these guidelines a strategy was recommended for the monitoring of metal-specific biological effects, incorporating a “cascade” of stages that included the measurement of metallothionein (MT), -amino levulinic acid dehydratase (ALA-D) and a range of “antioxidant enzymes”. The JAMP guidelines also drew attention to a number of “early warning” techniques, amongst which lysosomal stability was suggested as potential monitor of metal-mediated effects.

One problem affecting the application of the suggested biological effects techniques to sediments is that there is no commonly accepted chemical method of determining the biologically-available metal bound to sediment. Indeed, the proportion of bioavailable metals bound to sediment may vary depending upon the organism being studied. Since study of biological effects tends to rely upon some correlation with the metal concentration, the lack of an accepted sediment digestion technique is a serious drawback for these contaminants.

This document reviews the use of the JAMP cascade techniques for metals in research and monitoring programmes. In particular, it addresses their practicalities in a field-monitoring context and their application specifically to the investigation of metal-mediated biological effects.

2.2 : Lysosomal Stability

2.2.1 : Introduction

Lysosomes are key cell organelles that play a vital role in the catabolism of cellular components, intracellular transport and the sequestration of organic and inorganic pollutants and their metabolites. It is because of their involvement in the latter, i.e. as part of an organism's defensive mechanism against pollution, that they have attracted substantial interest as potential indicators of pollutant-induced stress.

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Evidence suggests that various forms of chemical and physical stress can damage or destabilise the lysosomal membrane, thus leading to the leakage of lysosomal enzymes into the cytosol. These enzymes, normally used for the breakdown of damaged internal structures or exogenous items taken in by phagocytosis, have the ability to damage cells. It is on this basis that the measurement of lysosomal stability has been given importance.

2.2.2 : Analytical Methods

The two main methods that have been deployed in marine monitoring of lysosomal perturbations focus on the stability of the lysosomal membrane. Both techniques measure membrane damage although in different ways. Comparative studies have shown that the two techniques correlate well (Lowe et al., 1995a).

The first, so called lysosomal latency, has been most widely used in fish tissue (Kohler, 1991; Kohler et al., 1992). Briefly, the technique requires the immediate removal of liver (or potentially other tissue) from the fish and its storage at -70 C. From this sample, upon which the assay is conducted, cryostat sections are cut at 25C. Then, the lysosomal membrane stability is assessed by measuring the time of acid labilisation required to fully destabilise the membrane. This is achieved by selecting a lysosomal enzyme marker, e.g. -N-acetylhexosamidase, and incubating the sections with a substrate for the enzyme marker, AS-BI-N-acetyl--D- glucosamidine. The point of full membrane destabilisation is when maximum staining in the lysosome occurs (using an appropriate dye), which denotes penetration of the membrane by the substrate. The faster this occurs the more damaged the membrane is considered to be.

The second widely used technique is the neutral red retention assay (NRR) (Lowe et al., 1992; Lowe and Pipe, 1994). The main differences between this technique and lysosomal latency is that live tissues are used and that NRR measures efflux from the lysosome and not influx. Briefly, the assay works on the basis that lysosomal contents are acidic in nature and therefore weak bases such as neutral red are rapidly taken up into them. Using a microscope, the point at which the dye ”leaks out” into the cytosol is determined by visual detection. The more rapid the leakage the more damaged the lysosomal membrane is considered to be.

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2.2.3 : Field Applications and Practicalities

The lysosomal latency technique has the advantage that samples can be taken, frozen and processed at a later date in the laboratory. The neutral red retention (NRR) assay, however, utilises live tissue and therefore the assay needs to be done in the field or the test specimens need to be transported back to the laboratory. The equipment for the NRR assay is very basic and therefore has scope for use in the field. Although the NRR assay is cheaper, a very significant drawback is the inability to control ambient temperature which can strongly influence the results.

The majority of organisms and tissues contain lysosomes, and therefore assays looking at lysosomal perturbations are particularly flexible with respect to the array of potential sentinel species. The lysosomal latency assay has been applied to clams (Nasci et al., 1999), but its most widely used application has been in the liver tissues of fish including flounder, Platichthys flesus (Kohler and Pluta, 1995), dab, Limanda limanda (Kohler et al., 1992) and plaice, Pleuronectes platessa (Kohler, 1991). It has also been extensively used in molluscs for which the mussel, Mytilus edulis, digestive gland is by far the most widely studied (Lowe et al., 1995a; Moore, 1990; Moore and Farrar, 1985; Widdows et al ., 1982).

The NRR assay requires a suspension of single cells and therefore has been most widely used in investigations into in bivalve molluscan haemocytes. The haemolymph is easily extracted with a syringe from the adductor muscle which means that the assay is the only one that can be classed as “non-lethal”. This technique has been most widely applied to the mussels Mytilus edulis (Lowe et al., 1995a) and Mytilus galloprovincialis (Lowe et al., 1995b) but has also been applied to the European flat oyster, Ostrea edulis (Hauton et al., 1998) and the freshwater gastropod, Viviparus contectus (Svendsen and Weeks, 1995). However, the technique can also be applied to other tissues by disaggregating the tissues with enzymatic treatments. This approach has been successfully applied to digestive gland tissue of mussels (Lowe and Pipe, 1994) and the liver of dab, Limanda limanda (Lowe et al., 1992). The disaggregation procedure inevitably adds an extra layer of difficulty for work in the field but is, in itself, relatively straight forward.

Field surveys using assays of lysosomal perturbation have been widely used. Pelletier et al. (1991) used lysosomal stability in mussels to investigate effects in oil spill contaminated sediment. The assay allowed differentiation between impacted and non-impacted areas. Other workers have performed interesting studies upon mussels, for example: Lowe et al. (1995b) investigated NRR in the haemocytes of mussels collected from a number of sites in the vicinity of Venice Lagoon, and Regoli (1992) used the lysosomal latency approach to investigate impacted areas in the Tyrrhenian Sea. The results demonstrated the ability of the assays to clearly differentiate between sites. Tissue analysis performed by Lowe et al. (1995b) indicated that effects could be attributed to a range of contaminants but, in particular cobalt, mercury and organochlorines. In Regoli's study (1992) the effects were also attributed to heavy metals, with elevated concentrations of copper, manganese, iron and lead at the contaminated site.

Kohler (1991) investigated lysosomal latency in the livers of flounder, Platichthys flesus, from three stations in the German Wadden Sea. This study showed a gradient

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of effect from the Elbe estuary which coincided with the concentrations of organochlorines and heavy metals in liver tissue. Perhaps the most useful examples of lysosomal perturbation monitoring in fish were the parallel studies of Lowe et al. (1992) and Kohler et al. (1992) in fish taken from a transect in the German Bight during the Bremerhaven workshop. The authors investigated NRR in disaggregated hepatocytes and lysosomal latency in liver tissue from the dab, Limanda limanda. The results of both studies were in close agreement, finding a seaward gradient of response including a marked reduction in lysosomal membrane stability at the outer station on the Dogger Bank. It could be argued that the NRR assays showed more pronounced differences between the sites suggesting that this technique has greater sensitivity. However, the sensitivity of NRR was so high at the inner stations that no differentiation could be observed between the sites, whereas the latency approach did differentiate between them. This suggests that lysosomal latency might be more applicable over a wider range of environmental concentrations. Lowe et al. (1992) suggested that at the inshore stations the organic compounds were the major contributors to the lysosomal damage, whereas metals were probably more important on the Dogger Bank. These studies, while demonstrating the use of lysosomal damage as a monitoring tool, also highlight its non-specificity and integrative characteristics.

2.2.4 : Influences on the Biomarker

The literature abounds with evidence that lysosomal membranes are affected by virtually all the major classes of contaminants. Laboratory studies have demonstrated the ability of PAH such as fluoranthene (Lowe and Pipe, 1994; Lowe et al., 1995a) and phenanthrene (Viarengo et al., 1987) and the heavy metals copper (Suresh and Mohandas, 1990; Svendsen and Weeks, 1995) and cadmium (Viarengo et al. 1987) to be able to compromise lysosomal membrane stability in molluscs. However, whilst chemical-induced stress has been shown to affect lysosomal integrity, there are a wide range of other influences that can affect the organelles and consequently the assays described above.

For example, the temperature at which the organism has been acclimated has been shown to have a very significant effect on lysosomal stability, especially in molluscs. Hauton et al. (1998) and Patel and Patel (1985) demonstrated this effect in the oyster, Ostrea edulis, and the tropical blood clam, Anadara granosa, respectively. In these studies, one using lysosomal latency and the other neutral red retention, disturbance in lysosomal stability was observed either side of an optimal temperature which appears to be species-specific. Furthermore, both studies noted a similar effect on lysosomal stability due to salinity. The reaction of the lysosomal system to external physical conditions such as these mean that responses to seasonal fluctuations, for example, would need to be fully established before the assays could be used confidently in monitoring programmes. Etxeberria et al. (1995) and Regoli (1992) noted pronounced seasonal fluctuations in lysosomal membrane stability of the mussel, Mytilus galloprovincialis. However, whereas Regoli (1992) suggested that this fluctuation was primarily due to thermal stress during the summer, Etxeberria et al. (1995) attributed this instead to other seasonal changes such as food availability.

The suggestion that feeding behaviour and food availability can affect lysosomal stability is plausible considering the pivotal role that lysosomes play in digestion

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(Etxeberria et al., 1995). Moore (1990) has highlighted dietary depletion as a probable cause of lysosomal alteration along with several other non-chemical stressors including hypoxia, hyperthermia and osmotic shock. Reproductive processes will also have an effect on lysosomal characteristics. The autophagic response in the digestive gland as reserves are switched to the gonads during gametogenesis, for example, will result in increased lysosomal activity (Regoli, 1992). The fluctuations in steroid hormone levels at this time will also be influential, as illustrated by the reduced lysosomal stability observed in mussels exposed to 17 -estradiol (Moore and Viarengo, 1987).

There is relatively little information in the literature regarding the effects of non-contaminant mediated lysosomal effects in fish, as compared to the relative plethora of information available for molluscs. However, their role in digestion, cell repair and contaminant sequestration is well documented and as such the stability of their membranes is subject to similar influences to those of invertebrates. Nevertheless, the greater ability of fish to move away from external stimuli and to control their internal environment may mean that their lysosomal stability is less variable. However, Kohler (1991) did note a pronounced seasonal difference in hepatic lysosomal stability in flounder, Platichthys flesus, associated with vitellogenin production during gonadal development.

2.2.5 : Discussion - Lysosomal stability

The two techniques described, lysosomal latency and neutral red retention, are both useful tools for measuring lysosomal stability, and comparative studies have shown good agreement between them. Therefore selection of the “preferred method” is more an issue of practicality than or sensitivity or accuracy. Where assay conditions are particularly difficult to control or the samples cannot be analysed soon after sampling, the lysosomal latency test would be selected because the samples can be stored. The advantage of the neutral red retention assay is that it is simple to perform (especially in the haemolymph of molluscs), and while it can be prone to analyst subjectivity it may be used to generate results quickly in the field.

The ubiquity of lysosomes across species and tissues means that this type of assay has the potential to be particularly flexible with respect to the range of situations in which it may be applied. However, as with all assays that are heavily influenced by non-contaminant stressors, significant study of lysosomal characteristics in a particular species needs to be performed before it could be applied with any confidence. Moreover although the majority of tissues contain lysosomes, there is a vast difference in the size and number of these depending on the cell function. In this respect the lysosomes of molluscan digestive gland tissue are particularly useful, whereas some other tissues such as fish hepatocytes have relatively few and smaller lysosomes under normal conditions (Kohler, 1991) making their visual assessment more difficult. Nevertheless, the lysosomal system of fish appears to afford greater stability with respect to physiological changes during reproduction and physical stress than is documented for molluscs (Kohler et al., 1992). The choice of assay, species and tissue is not straightforward and must be done after consideration of the aims of the survey. This in-depth investigation has been carried out for certain species and lysosomal stability in fish and Mytilus spp. has been recommended by ICES at the national/international level as a measure of cellular damage (ICES, 1999).

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It is clear that measurements of lysosomal stability can be used as valuable monitoring tools of contaminant exposure. They respond to a wide array of contaminants of which metals are a significant group. As a result of this non-specificity, the use of such techniques specifically for the monitoring of metal pollution can only be limited, but is perhaps worth consideration in areas where it is known that metals are the primary contaminant. Since lysosomal assays are fairly non-specific with respect to particular contaminants, they should be viewed as a general indicator of deterioration in the health of the target organism (Moore, 1990). Furthermore, as with other assays using biological systems that can be affected by non-contaminant stressors and that interact significantly with other biological systems, they are best used as part of a battery of biomarkers after careful consideration of the test species and external influences.

2.3 : Oxidative Stress

2.3.1 : Introduction

The presence of highly reactive oxygen species (oxyradicals) such as the superoxide anion radical (O2

-) and hydrogen peroxide (H2O2) in tissues can be detrimental at many levels. The effects they produce can include; changes in redox balance, general cellular and skeletal damage, lipid peroxidation and DNA damage (Livingstone et al. 1992). Moreover, increased concentrations of oxyradicals have been shown to occur as a response to contaminant exposure, so assays for oxyradicals and associated enzymes (produced as a natural response of tissues to neutralise the presence of reactive oxygen species) have been suggested as monitoring tools.

Several biomarkers have been proposed including; superoxide dismutase (SOD) which converts O2

- to H2O2; catalase which converts H2O2 to water: glutathione peroxidase (GPX) which also converts H2O2 to water; glutathione reductase (GR) which maintains cellular reduced glutathione (an important antioxidant); and malonedialdehyde which is a measure of lipid peroxidation.

2.3.2 : Analytical Methods

Various methodologies exist for the measurement of oxidative stress biomarkers. The majority are relatively simple spectrophotometric assays. As a general rule the biomarkers are measured in relatively crude supernatants or require uncomplicated extraction/filtration procedures, and the techniques do not pose significant barriers to their application in the field. Furthermore, tissues can be cryogenically stored for analysis at a later date if required.

2.3.3 : Field Application and Practicalities

Superoxide dismutase (SOD) has been measured in a variety of species including fish such as the dab (Livingstone et al., 1992), lake trout (Palace et al., 1998), in molluscs such as mussels (Sole et al., 1995) and even in cephalopods (Zielinski and Portner, 2000). Whilst its use in the field is relatively straight forward, the results have been ambiguous. Livingstone et al. (1992) tentatively showed an increase in SOD activity

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in dab at the contaminated coastal stations during the Bremerhaven workshop, whereas van der Oost et al. (1996) found no site- or pollutant-related response to SOD in eel. Nasci et al. (1998) also found the pattern of SOD response in mussels difficult to relate to contamination.

Catalase (CAT) has been widely investigated across a broad range of species including dab (Livingstone et al., 1992) and mussels (Regoli and Principato, 1995). Again the field data are inconclusive, e.g. van der Oost et al. (1996) found highest CAT activity in eel at their reference site, and they stated that this was inconsistent with other studies which have observed elevated activity correlated to contamination (Rodrigues-Ariza et al., 1993).

The activity of glutathione reductase (GR) and glutathione peroxidase (GPx) are also measurable in most phyla. Doyotte et al. (1997) used caged bivalves, Unio tumidis, to investigate the effects of a cokery discharge. While other antioxidant enzymes (SOD, CAT and total GPx) showed no change, selenium dependent GPx and GR showed marked decreases in activity in exposed specimens. Conversely, van der Oost et al. (1996) concluded that GPx was not a good biomarker for contaminant effects in the eel.

Lipid peroxidation (LP) has been shown to increase in dab after exposure to contaminated sediments (Livingstone et al., 1993). For molluscs, Sole et al. (1996) noted an increase in LP at a site near to the "Aegean Sea" oil tanker spill in the mussel, Mytilus edulis. However, Doyotte et al. (1997) and Pellerin-Massicote (1994) showed no increase in LP in a freshwater bivalve (Unio tumidus) or Mytilus edulis respectively in field surveys of contaminated areas.

In summary, these biomarkers are practical to use, as is borne out by their wide application to a range of species, but the field data highlight the difficulties in obtaining useful monitoring data that are well correlated to contaminants.

2.3.4 : Influences on the Biomarkers

A recurring theme in the literature is the variability in response of biomarkers of oxidative stress. Laboratory data confirm that the systems triggered by oxidative stress respond to contaminant exposure but field data routinely provide ambiguous results. One reason for this is the fact that as a class, the biomarkers of oxidative stress are highly influenced by a range of factors other than xenobiotic exposure. For example, GPx activity has been shown to be age dependent in carp (Machala et al., 1997) and age dependence has also been demonstrated in cephalopods for SOD, GPx and CAT (Zielinski and Portner, 2000). McFarland et al. (1999) also noted significant inter-gender differences in activity of GPx and GR in the brown bullhead, Ameirus nebulosus.

SOD activity has been shown to demonstrate wide seasonal variation in a fish (Palace et al., 1998) and mussels (Sole et al., 1995). Pellerin-Massicote (1994) showed a similar seasonal variation for lipid peroxidation in the mussel, Mytilus edulis. As would be expected, another significant factor is the level of anoxia to which the target organism has been exposed. Pannunzio and Storey (1998) investigated different conditions of anoxia exposure and aerobic recovery upon the gastropod, Littorina

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littorea, for SOD, CAT, GR and GPx. Aerobic recovery may be of particular relevance to the use of these assays in areas of fluctuating oxygen levels e.g. estuaries and intertidal zones.

Other important considerations involve the choice of tissue in which the analysis is performed. Thomas and Wofford (1993) studied GPx activity in contaminant-exposed fish (Atlantic croaker), and found that GPx activity was higher in ovary than in liver, and that the inhibition in activity associated with exposure was only significant in ovarian tissue. An even more complicated tissue-dependent picture was shown by Regoli and Principato (1995) on exposing the mussel, Mytilus galloprovincialis, to copper. In this instance the responses of CAT and GPx in digestive gland were stable or decreased, but they were increased in gills.

2.3.5 : Discussion - Oxidative stress

The biomarkers of oxidative stress reported above have been shown to exhibit a concentration response in several taxonomic groups following exposure to heavy metals. Examples are the increased levels of lipid peroxidation in mussels following exposure to copper (though Cd and Zn showed no effect) (Viarengo et al., 1990); fish (Atlantic croaker) following exposure to cadmium (Thomas and Wofford, 1993); and a freshwater crab following exposure to chromium (Sridevi et al., 1998).

However, while these techniques are flexible and straight-forward to perform, their suitability for monitoring purposes for metals is questionable because they are influenced by a range of other factors. Furthermore, several other classes of chemicals have been shown to elicit responses in antioxidant enzymes (e.g. PAH and PCBs) so their specificity to metals is very poor. Nevertheless, as with the other biomarkers these measures of oxidative stress may have a place as part of a battery of techniques for monitoring environmental quality. The ICES Working Group on Biological Effects of Contaminants (WGBEC) felt there were sufficient data to recommend application of oxidative stress techniques in fish (ICES, 1999).

2.4 : Metallothionein

2.4.1 : Introduction

Metallothioneins are low molecular weight metal-binding proteins. They are present in most tissues of most phyla. They play a vital role in the homeostatic control of the essential trace elements zinc and copper, and act as part of the detoxification mechanism for other metallic pollutants, especially cadmium and mercury. Metallothionein (MT) synthesis above baseline levels can be induced in a wide selection of species as a response to exposure to exogenous metals. MT has therefore been suggested as a biomarker for metal contamination.

2.4.2 : Analytical Methods

There are many techniques in the literature that have been used for the determination of metallothionein in aquatic species. The three main methods suggested in the JAMP Guidelines for Contaminant-specific Biological Effects Monitoring are i)

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immunochemical techniques (ELISA/RIA) (Hogstrand and Haux, 1990a); ii) differential pulse polarography (DPP) (Olafson and Sim, 1979); and iii) spectrophotmetric assay (Viarengo et al., 1997).

Of these methods the immunochemical techniques offer the greatest potential with respect to accuracy and specificity. In general, enzyme-linked immuno-sorbent assays (ELISAs) and radio immuno-assays (RIAs) tend to be highly specific and can demonstrate poor reactivity between species. However, for metallothionein (MT) the amino acid sequences appear to have been conserved between different fish species (Roesijadi, 1992), which suggests that similar antisera could be used for a variety of species. For instance, antisera raised against perch, Perca fluviatilis, (Hogstrand and Haux, 1990a) has shown good cross-reactivity with dab, Limanda limanda, (Hylland et al., 1992), flounder, Platichthys flesus, (Goksoyr et al., 1996), rainbow trout, Oncorhynchus mykiss, (Hogstrand and Haux, 1990a) and subtropical fish species (Hogstrand and Haux, 1990b).

The differential pulse polarography technique has been widely used to measure MT, for example, Stagg et al. (1992) applied it to dab; Temara et al. (1997) to echinoderms; Olsson and Haux (1986) to perch. DPP has been described as allowing rapid and accurate determination of MT (Olsson and Haux, 1986), as being considerably less difficult than other techniques (Thompson and Cosson, 1984) and has been shown to have a close linear relationship with a RIA method (Hogstrand and Haux, 1990a). However, Romeo et al. (1997) stated that the polarography method might tend to overestimate MT levels, because of the possibility of disturbance by sulphur proteins if these were not efficiently removed during the denaturation/centrifugation process. In their study, measurements of MT by DPP were higher than those made spectrophotometrically (ibid).

A simple spectrophotometric assay was suggested by Viarengo et al. (1987). However, few other authors have reported using this method, and whilst Viarengo et al. (1997) used it successfully for the determination of MT in molluscs, the only other studies in the literature were by Romeo et al. (1997) and Pedersen et al. (1997). Romeo et al. (1997) performed a comparison between a spectrophotometric assay and the DPP technique for the determination of MT in the liver of the bass, Dicentrarchus labrax. While the spectrophotometric technique was able to differentiate between MT levels in various treatments, it recorded lower MT concentrations than did the DPP method. Pedersen et al. (1997) used the spectrophotometric method to measure MT in Carcinus maenas. They found that MT in the gill material reflected metal exposure gradients.

Another category of widely used techniques, not included in the JAMP suite, is that of metal substitution. These techniques rely on the fact that metals have differential binding affinity with MT, so that those for which MT has a stronger affinity will displace those for which it has a lower affinity. Metal binding strength with MT increases in the order zinc < cadmium < copper < mercury < silver (Hamilton and Mehrle, 1986). The most widely used technique appears to be that of cadmium saturation (Overnell et al., 1987; Sulaiman et al., 1991), but this method could result in the underestimation of total MT as Cd will not displace all bound metal. Silver, because of its higher binding affinity with MT, has also been used (Klaverkamp et al., 1996). The main problem with metal binding techniques is that that they are an

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indirect measurement of MT, and the binding mechanisms under different conditions and for different biological species do not appear to be fully understood. Chan (1995) points out that metal-saturation techniques can be inaccurate in their determination of MT content because other non-MT proteins might also bind metals. They may also underestimate MT concentrations as these techniques do not provide information upon metal ions bound to MT in vivo.

More recently, the measurement of MT mRNA has been suggested as a highly specific and useful biomarker of MT induction (Kille et al., 1992). As with ELISA and RIA techniques, mRNA offers the opportunity to measure the very specific onset of MT synthesis. Methods also appear to offer cross-reactivity with a number of piscine species. However, the significance of this measurement is questionable from a monitoring viewpoint, as it is the MT protein itself that is biologically relevant and MT mRNA may not accurately reflect this as its half-life is much shorter than that of MT. MT concentrations may remain elevated long after the mRNA levels have dropped (Romeo et al., 1997).

There is clearly a need for the methodology to be standardised and for appropriate quality assurance procedures to be established. This is currently being addressed through the European Union funded BEQUALM programme (Biological Effects Quality Assurance in Monitoring Programmes). The purpose of this project is to develop quality assurance and control procedures for marine biological effects measurements, in order that laboratories contributing to international marine monitoring programmes such as the OSPAR Joint Assessment and Monitoring Programme (JAMP) can attain defined quality standards. Metallothionein is among the techniques being assessed under the programme, together with lysosomal stability, ALA-D and EROD, for example.

2.4.3 : Field Applications and Practicalities

A particular advantage of studying MT is that its relative stability means that samples can be taken and stored in liquid nitrogen for analysis at a later date. This means that the choice of technique does not impinge on the practicality of MT measurement in field samples. Analysts wishing to measure MT in the field itself can select the simpler techniques such as spectrophotometric assay in preference to the more involved methods e.g. immunochemical techniques.

Another advantage is that MT has been reported and measured in many species. Roesijadi (1992) lists 75 different aquatic species including fish, echinoderms, insects, crustacea, bivalve and gastropod molluscs, oligochaetes and polychaetes in which MT has been measured. The wide inter-species occurrence of MT means that it has the potential to be applied as a monitoring tool in many different geographical areas and habitats by selection of an appropriate species.

MT determinations have been applied in many field studies. Correlations were found between hepatic MT levels and hepatic Cd concentrations in perch (Perca fluviatilis) from a contaminated river in Sweden (Olsson and Haux, 1986), and a similar relationship was found between MT and zinc in rainbow trout in a Canadian river study (Roch and McCarter, 1984). Klaverkamp et al. (1996) measured MT in various tissues in lake whitefish and pike from a metal contaminated lake in Canada. They

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showed some correlation with metal concentrations in tissue but a confusing picture with respect to links with metal concentrations in sediments.

Of more significance to the JAMP are the field studies that have been conducted in estuarine and marine environments. Sulaiman et al. (1991) measured hepatic metallothionein in flounder, Platichthys flesus, at several sites along the Forth estuary. They discovered high variability of levels within each sample but showed a general trend of MT levels increasing in lower/mid region of the estuary. Furthermore there was some indication that these levels corresponded with hepatic metal concentrations (Cu, Cd, Zn and Pb). Pedersen et al. (1997) found that copper and zinc in gill material from crabs in clean and polluted estuaries reflected the sediment concentrations, and also that the MT reflected exposure gradients.

Of great interest were the parallel studies of Hylland et al. (1992) and Stagg et al. (1992) who investigated MT levels in dab, Limanda limanda, caught along the Bremerhaven transect. Hylland et al. (1992) found that hepatic MT concentrations were highest at stations offshore compared to those nearer the shore. This conflicted with the sediment trace metal concentrations which decreased away from the shore, but did to a certain extent reflect the hepatic accumulation of Cu, Zn and Cd. Differences between the sexes were observed, with female hepatic MT levels demonstrating a correlation between Zn and MT (r=0.48), while that in males was stronger with Cu and Cd, with weak but significant correlations of r= 0.29 (p=0.017) and r=0.27 (p=0.025). The authors speculated that the trends could have been due to the variable metal concentration in the diet, as Cofino et al. (1992) had shown that certain benthic invertebrates had higher trace metal concentrations offshore compared to those found nearer the coast. Conversely, Stagg et al. (1992) found that MT concentrations in the gills of dab decreased away from the shore in a similar fashion to those of sediment metal concentrations. These authors also found a relationship between MT levels in gills and the sum of the molar concentrations of Zn, Cd and Cu in the same tissue. These two studies, while demonstrating the potential use of MT as a biomarker for metal exposure in the marine environment, also show that there is not a simple relationship between the two and that the data may need much careful interpretation.

Field studies in the marine environment have not been limited to the use of fish. Viarengo et al. (1999) used MT in the mussel to identify heavily contaminated areas in the Mediterranean, and Galgani et al. (1992) measured it in oysters from the Gironde estuary. In the latter study no relation was found between MT concentrations and the very high Cd concentrations associated with the area. Pedersen et al. (1997) investigated MT measurements in the shore crab, Carcinus maenas, and Temara et al. (1997) have investigated MT in the common asteroid starfish, Asterias rubens. Both groups have had some success with respect to differentiating between sites but not necessarily with establishing cause-effect relationships.

2.4.4 : Influences on the Biomarker

The literature shows that MT is induced in a wide variety of species as a result of heavy metal exposure (specifically Zn, Cd, Cu and Hg). This has led several authors to endorse the use of MT as a biomarker for metal exposure in the field. However, as

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can be seen from the selection of studies above, the relationship between MT levels and metal concentrations in the environment is not simple.

A general review of the literature reveals that MT levels can be influenced by a range of factors other than metal exposure, and that there is contradictory evidence for certain effects. For example, stress from netting and/or capture has been reported as an inducer of MT in fish (Romeo et al., 1997). It is thought that capture stress raises cortisol levels which has been shown to induce MT (Baer and Thomas, 1990). However, Overnell and McIntosh (1989) stated that “stress factors” are not important inducers in marine flatfish, and earlier work by Overnell et al. (1987) failed to induce MT in plaice, Pleuronectes platessa, after cortisol injection.

Physical conditions are also thought to affect MT levels. Tom et al. (1999) found higher levels of MT mRNA induction in a sparid fish kept at a lower temperature (21ºC as opposed to 27ºC). Viarengo et al. (1999) noted that temperature, salinity and oxygen content could all influence MT levels in the mussel, Mytilus galloprovincialis, although it was noted that heavy metals were far more influential in determining the MT level. Legras et al. (2000) demonstrated a link between salinity and MT levels in crabs, which might be expected when one considers the role salinity can play in metal bioavailability and uptake.

The issue of seasonal trends is very important in determining the baseline of MT in several, if not most, species and clear gender differences can be apparent. In dab, for example, Hylland et al. (1992) found slightly different trends between each sex for MT induction in liver tissue along the Bremerhaven transect. Although there was some agreement between the sexes, female MT levels were associated with hepatic Zn concentrations whereas in males they were slightly more related to Cu and Cd. The authors suggested that the differences were possibly triggered by physiological processes related to sexual maturation and a spawning. This hypothesis seems reasonable, as Baer and Thomas (1990) noted a two-fold increase in Zn binding during ovarian maturation and significantly higher Zn binding in ripe individuals compared to earlier stages.

A clear seasonal trend in digestive gland tissue MT was evident in the mussel, Mytilus galloprovincialis, with concentrations in mid-summer two-fold higher than in the winter months (Viarengo et al., 1997). This, as in fish, appears to be associated with a rise in normal levels of zinc. MT has been proven to play a major role in the regulation of Zn and Cu levels during ecdysis in crustacea (Engel and Brouwer, 1991), which can lead to very sharp fluctuations in MT. While Pedersen et al. (1997) selected the shore crab, Carcinus maenas, as their biomonitor species, they acknowledged the drawbacks of this species that MT is involved in haemocyanin synthesis and metal handling during the moult cycle. The increase of MT during ecdysis has been experimentally linked to the activity of the moult regulating hormone hydroxyecdysone by Torreblanca et al. (1996) who noted a marked increase in MT content of the crayfish, Procambarus clarkii, after its injection.

When using any biomarker system in a biomonitoring context one must also consider possible interferences. For example, other enzymatic and endocrine systems could interact with the specific biomarker system, and other contaminants could influence the level of this interaction. The example of hydroxyecdysone above demonstrated

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the potential hormonal influences on MT levels. Glucocorticoids and a number of other hormones, for example oestradiol, have been shown to regulate MT at the cellular level in fish (Baer and Thomas, 1990; Olsson et al., 1989). These are associated with the fluctuations seen during the breeding cycle and the onset of vitellogenesis, but also suggest that MT levels may be under some indirect influence from a range of endocrine disrupters known to be present in UK rivers/estuaries (Matthiessen et al., 1998). Furthermore, an association has been made between MT levels and the activity of the mixed function oxygenase (MFO) system (Romeo et al., 1997). The MFO enzyme system is induced as a detoxifying response following exposure to certain planar organic compounds (especially some PAH and PCBs), and its action can stimulate the production of reactive oxygen species for which MT is thought to be a scavenger, resulting in enhanced synthesis of MT. In contrast, the same authors suggested a mechanism whereby MT synthesis was could be decreased, owing to the demands for cysteine residues (which are high in MT) for glutathione production during organic contaminant detoxification. The potential for these MT/MFO interactions are high in industrial estuaries where the MFO system has been shown to be highly induced (Kirby et al., 1999).

The main advantages of a biomonitoring technique based around metallothionein are therefore that the protein is found in a wide range of species and it does appear to be induced by metal exposure. However, any use of MT in a monitoring context would have to take great care in choice of organism and particular tissue. Such an approach would also have to avoid the conflicting physical and seasonal factors described above. Hogstrand and Haux (1990b) noted high levels of MT, Zn and Cu in the squirrelfish, Holocentrus rufus, even at control sites that appeared to be unaffected by metal contamination. There is therefore a need to investigate baseline MT levels and seasonal influences before selecting a species for biomonitoring purposes. Furthermore, there is much evidence in the literature which indicates that tissue selection is an important issue. Klaverkamp et al. (1996) demonstrated significantly different MT levels in liver, kidney and gill of the same fish that could result in differing interpretation of field data. Temara et al. (1997) highlighted a similar situation in echinoids, where tissue from the gonad or pyloric caeca are shown to react differently. Perhaps the clearest example of this is in the comparison of gill and liver MT levels in dab caught along the Bremerhaven transect which resulted in quite different interpretations of the results (Stagg et al. 1992; Hylland et al. 1992).

2.4.5 : Discussion - Metallothionein

Selection of the preferred analytical technique is very difficult as there are few direct comparisons between them in the literature. However, the three techniques listed in the JAMP Guidelines (immunochemical, differential pulse polarography and spectrophotometry) for Contaminant-specific Biological Effects Monitoring (OSPAR 1998) appear to be satisfactory, as they have been extensively applied and have generally been shown to correlate well with each other. Identifying the optimal method(s) is further hampered since authors tend to favour their own techniques. The need to directly compare analytical techniques under the same conditions and using the same species is being addressed under the EC-funded BEQUALM programme (Biological Effects Quality Assurance in Monitoring Programmes).

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Metallothionein can be induced in a wide variety of species as a direct result of exposure to particular heavy metal contaminants, and in certain cases this has lead to the successful application of the biomarker in monitoring (eg MEDPOL). However, it is evident that in a number of other studies the technique has failed to clearly highlight contaminated areas; furthermore there is conflicting evidence regarding the significance and influence of the numerous factors that can affect MT levels. It is apparent that MT cannot be reliably used to monitor for the presence of metal contamination in the environment per se, but its induction is probably representative of the bioavailability of the metals and their accumulation in tissues. From a biological monitoring perspective, this is a more relevant measure.

Metallothionein remains one of the few, and probably the best, biomarker for metal exposure, and as is a potentially important tool for the ecotoxicologist. ICES WGBEC recommended its application in fish for national programmes (ICES, 1999). However, because of the various factors that can influence its expression, it can only be useful if the seasonal dynamics of MT are well understood in the chosen species and in the selected tissues. Furthermore, as with many other biomarkers, the data provided by MT are of most benefit if the test is deployed as part of a wide suite of assays (e.g. EROD) and chemical analyses (specifically heavy metals), so that the results can be interpreted with consideration to other relevant factors.

2.5 : -Amino Levulinic Acid Dehydratase (ALA-D)

2.5.1 : Introduction

The enzyme -aminolevulinic acid dehydratase (ALA-D) plays a pivotal role in the production of haemoglobin and other porphyrin-based structures in fish and other vertebrates. ALA-D catalyses the reaction that results in porphobilinogen (PBG) being formed from two molecules of -aminolevulinic acid (ALA). This reaction is an early stage in the production pathway of haem and consequently the maintenance of haemoglobin. A significant and prolonged inhibition in the activity of ALA-D could therefore result in anaemia and detrimental effects upon oxygen transport mechanisms.

Certain heavy metals, principally lead, have been found to act as efficient inhibitors of ALA-D activity in fish (Hodson, 1976; Krajnovic-Ozretic and Ozretic, 1980), and therefore this assay has been suggested as a biomarker for lead exposure.

2.5.2 : Analytical Methods

There does not appear to be a “standard” methodology for the ALA-D assay. However, all the methods investigated as part of this literature survey were based on a straight forward spectrophotometric procedure. Briefly, the tissue is homogenised (or haemolysed for blood cells) and used as raw material for the source of ALA-D. This material is incubated with ALA as the substrate, and the activity is based on the rate of synthesis of the product (PBG). The product is reacted with a colour producing reagent (normally Ehrlich's), and the absorbance read spectrophotometrically. The activity is finally expressed as a factor of weight of tissue, volume of blood/tissue, mg protein or mg haemoglobin (in blood). Again, no standardisation is apparent.

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There is some evidence that ALA-D has been successfully determined in crustacean species, for example the cladoceran, Daphnia magna (Berglind et al., 1985; Berglind, 1986), but Brock and Brock (1993) found no ALA-D activity in the hepatopancreas of the bivalve mollusc, Cerastoderma edule. This suggests that this assay may not be suitable for molluscan species. However, the majority of the literature is based on the use of both marine and freshwater fish species such as rainbow trout, Oncorhynchus mykiss (Burden et al., 1998), channel catfish, Ictalurus punctatus (Conner and Fowler, 1994), carp, Cyprinus carpio (Nakagawa et al., 1995) and flounder, Platichthys flesus (Johansson-Sjobeck and Larsson, 1978). ALA-D can also be measured in a wide range of tissues including liver, spleen and kidney, but has been predominantly measured in erythrocytes isolated from the blood.

2.5.3 : Field Application and Practicalities

The ALA-D assay has been applied to fish from a wide range of environments. Moreover, since blood is frequently used, there is the opportunity for non-destructive sampling. In cases of small species or juveniles, where enough blood cannot be extracted or specific tissue excision is impractical, ALA-D has successfully been used to monitor lead poisoning in whole body homogenates (Burden et al., 1998).

The assay is relatively easy and inexpensive to perform, and in general could be performed in the field given access to a spectrophotometer. Moreover the enzyme is relatively stable during storage. Hodson et al. (1984) stated that blood samples could be held on ice for up to 24 hours without loss of activity (as long as they were heparinised to stop clotting) and deep freezing would maintain activity indefinitely.

The studies of Schmitt et al. (1984) and Dwyer et al. (1988) investigated the ALA-D levels in blood of fish caught in the Big River, Missouri, USA - an area contaminated with mine tailings containing high concentrations of heavy metals. Schmitt et al. (1984) sampled 3 catostomid fish (the black redhorse, golden redhorse and northern hogsucker) whilst the Dwyer et al. (1988) investigated effects in the longear sunfish, Lepomis megalotis. Both studies demonstrated that fish caught downstream of the contaminant sources had significantly elevated erythrocytic ALA-D levels when compared to a reference site upstream, and this was correlated to blood lead levels. Although the studies did not detect any obvious detrimental consequences of lead exposure in the fish, Dwyer et al. (1988) suggested that certain bone variables, such as strength, were probably affected by prolonged lead exposure.

Martin and Black (1998) used blood ALA-D levels in caged channel catfish, Ictalurus punctatus, to investigate heavy metal induced effects in contaminated ponds at an abandoned coal strip mine. In this study the ALA-D levels at all sites were highly variable and significant differences as compared to a reference site were only apparent on one sampling occasion. The authors suggested that the assay can suffer severe interference when significant concentrations of metals other than lead are present. A study by Hodson et al. (1984) measured ALA-D in fish caught near an alkyllead producer on the St Lawrence River. They also noted the high variability of ALA-D measurements. Although a trend related to lead exposure was apparent, it did not reflect the large differences of contamination between the sites. The authors suggested that most of the lead was alkylated - a form in which it does not strongly inhibit

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ALA-D. In a BEQUALM study, significant differences were not detected between ALA-D measured in lead-exposed-cod compared to non-exposed cod (ICES, 2001). In the same study, unexposed flounder showed higher ALA-D induction than exposed flounder, so further method development appears necessary.

It is worth mentioning a study by Haux et al. (1986) who looked at blood ALA-D levels in whitefish (Coregonus spp.) living in lead-contaminated lakes in Sweden. In this case a very clear inhibition of ALA-D was noted at the contaminated sites, but again no evidence of adverse physiological effects was apparent.

It is clear that ALA-D is a sensitive biomarker of lead exposure and in some cases it can be successfully used to investigate exposure in fish populations. However, whilst the assay is suited to field studies from a practical viewpoint, it is difficult to draw firm conclusions as to its utility as a monitoring tool because of the apparent paucity of data concerning its use in the marine environment and on non-fish species. This may be attributed, in part, to the low number of marine sites where lead contamination is a significant problem.

2.5.4 : Influences on the Biomarker

With the ALA-D assay, as with all enzyme-based markers, the results are critically dependent upon assay conditions. The lack of method standardisation across the literature makes this particularly important for the ALA-D test. Conner and Fowler (1994) demonstrated that incubation temperature and pH can produce large differences in the assay result for hepatic ALA-D of the channel catfish, Ictalurus punctatus. Krajnovic-Ozretic and Ozretic (1980) showed a similar effect in samples of blood from the grey mullet, Mugil auratus, and also highlighted the reduction in ALA-D activity when substrate (ALA) concentration is not optimised.

There is some indication in the literature that ALA-D varies as a factor of size/age of certain fish species. Dwyer et al. (1988) showed that a range of variables, including tissue metal concentrations and skeletal biochemical properties (phosphorus, collagen, proline content etc.), were correlated to age/size and this was reflected in a moderate correlation (r2 = 0.57) of ALA-D with weight in longear sunfish. Furthermore, Hodson et al. (1984) stated that fish size affects lead accumulation and Burden et al. (1998) suggested that juvenile fish are more sensitive to lead. While inconclusive, these data suggest that age/size must be carefully controlled in any monitoring applications. Capture stress is another factor that can have potential effects on the biomarker, as Haux and Sjobeck (1985) noted an increase in blood ALA-D activity in perch (Perca fluviatilis) that remained elevated for 2-3 days after capture.

Seasonal fluctuations associated with the reproductive cycle are also likely to occur. Larsson et al. (1985) highlighted an example using perch for which erythrocytic ALA-D decreased during the summer spawning period. The mechanisms of interference appear to be poorly understood, but ALA-D is an integral component of a system that could be heavily influenced at the endocrine level and therefore seasonal fluctuations should be investigated for any candidate monitoring species.

The poorly understood influence of other metals on enzyme activity is one of the most significant difficulties in the interpretation of ALA-D data. It is generally agreed that

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lead is by far the most significant metal with respect to its inhibitory effects on the ALA-D enzyme in fish. However, while Hodson et al. (1984) suggested that near-lethal levels of Cu, Hg, Cd and Zn had no effect on ALA-D, more generally the literature provides conflicting evidence as to the ability of other metals to interfere and contribute to ALA-D activity levels. Cadmium exposure, for example, has been shown to stimulate ALA-D activity in some studies (Berglind, 1986; Jackim, 1973; Johansson-Sjobeck and Larsson, 1978), whereas others have demonstrated its ability to inhibit the enzyme (Rodrigues et al., 1989; Krajnovic-Ozretic and Ozretic, 1980), although the latter studies were in vitro exposures. Zinc has also been shown to have stimulatory effects on ALA-D (Jackim, 1973; Krajnovic-Ozretic and Ozretic, 1980) whilst copper and mercury are net inhibitors of activity (Jackim, 1973; Krajnovic-Ozretic and Ozretic, 1980; Rodrigues et al., 1989). Other compounds can affect the enzyme and confuse the issue, for example Bengtsson et al. (1988) demonstrated that tetrachloro-1,2-benzoquinone (TCQ), a component of bleached kraft mill effluents, has the ability to stimulate ALA-D activity in fourhorn sculpin (Myoxocephalus quadricornis).

2.5.5 : Discussion - ALA-D

ALA-D is regarded as being specific for lead, although there may be some degree of interference in the presence of other metals. While the specificity of ALA-D inhibition to lead means that the assay can be very targeted and useful for certain applications, it also means that these applications are limited to lead-contaminated areas. Furthermore, although ALA-D activity of rainbow trout is inhibited at waterborne lead concentrations as low as 5 g/l, actual concentrations in UK marine waters do not appear to exceed 0.2 g/l (CEFAS 1998). Lead has a very low solubility in sea water, so application of the assay in marine waters is of limited value. However, the assay may be more useful in marine species that are associated with lead contaminated-sediments.

At a practical level, the assay is inexpensive, relatively easy to perform and the enzyme activity is conserved well during storage. The literature suggests that it has only been applied with any success to fish species, providing an index of exposure which can be used in monitoring programmes (ICES, 1999). Evidence suggests that assay conditions and physiological condition of the specimens can significantly affect the results. Therefore, the apparent lack of a standardised approach throughout the literature implies that more work would need to be focused here before it could be confidently used in monitoring, especially if data comparisons between laboratories are required. This is borne out by the results of a recent BEQUALM inter-comparison where samples from exposed fish did not show the expected differences in comparison with non-exposed fish.

ALA-D is probably the most target-specific biomarker currently available and, together with metallothionein, one of the very few that are specific for metals. However, since lead has low bioavailability at the pH found in seawater, concentrations of the metal are generally below the threshold required for effects upon ALA-D to be observed. The presence of other metals can influence the result, which means any use of the assay should consider other metals that may be present. ALA-D does however, remain a very useful tool in the investigation of areas where Pb contamination is known or suspected to be an issue. As with other enzyme markers it

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is best used as part of a suite of other assays and in species whose physiology is well understood.

2.6 : Conclusions and Recommendations

Many biomarker measurements now have a proven track record in showing early changes to biological systems when organisms are exposed to contaminants. This literature review has considered the biomarkers lysosomal stability, oxidative stress, metallothionein (MT) and ALAD, and specifically their application in the field.

Conclusions

Each of the biomarkers (lysosomal stability, oxidative stress, MT and ALA-D) shows a response to metal exposure. Current dissolved metal concentrations are unlikely to reach high enough levels in UK marine waters to result in detectable effects using these assays. However, in several UK estuaries, metal concentrations in sediments, or bioconcentration or bioaccumulation effects, could produce detectable changes in these biomarkers.

The techniques vary in their specificity to metals. Lysosomal and oxidative stress approaches are indicators of general stress. MT is regarded as specific for metals (primarily cadmium, copper, mercury and zinc) and ALA-D is regarded as being specific to lead.

All of the biomarkers are affected by a range of physical and seasonal influences that can make the interpretation of results difficult to correlate with metal concentration. More work is required to understand these cycles.

Techniques measuring oxidative stress and ALA-D are probably the most straight-forward to perform. Lysosomal stability (especially the NRR technique) is rather subjective and its application requires highly skilled operators.

Expert groups recommend that currently, biomarker measurements should be applied comparatively in a hypothesis-driven manner (ICES, 2000).

Recommendations

Lysosomal stability and oxidative stress show general indications of stressed populations, rather than any contaminant-specific effects. Lysosomal perturbations in Mytilus spp. and fish have been recommended for monitoring a wide range of xenobiotic contaminants and metals. Oxidative stress indicators in fish have been recommended for monitoring environmental contaminants.

Metallothionein induction provides a method for the monitoring of metal exposure in a wide range of organisms, although detailed work is required to develop protocols. ALA-D can be useful in specific investigations into lead in fish, although its application in marine waters is limited.

More information is needed on how a range of species react to metal contamination with respect to all the biomarkers. Biomarker response in different tissues is an area requiring further research.

More information is needed on the physical (temperature, salinity etc.) and seasonal effects upon the biomarkers, to allow more confident interpretation of data from environmental samples.

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None of the biomarkers discussed can offer particularly useful monitoring data when applied in isolation. Where possible they should be used in conjunction with a battery of other techniques (e.g. chemical, ecotoxicological) to enable a more complete picture of the health of a population to be established.

2.7 : References

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Dwyer, F.J., Schmitt, C.J., Finger, S.E. and Mehrle, P.M. (1988). Biochemical changes in longear sunfish, Lepomis megalotis, associated with lead, cadmium and zinc from mine tailings. Journal of Fish Biology, 33, 307-317.

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Matthiessen, P., Allen, Y.T., Allchin, C.R., Feist, S.W., Kirby, M.F., Law, R.J., Scott, A.P., Thain, J.E., Thomas, K.V. Estrogenic endocrine disruption in flounder (Platichthys flessus L.) from United Kingdom estuarine and marine waters (1998) Science Series Technical Report No. 107, 48pp CEFAS

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Schmitt, C.J., Dwyer, F.J. and Finger, S.E. (1984). Bioavailability of Pb and Zn from mine tailings as indicated by erythrocyte -aminolevulinic acid dehydratase (ALA-D) activity in suckers (Pisces: Catostomidae). Canadian Journal of Fisheries and Aquatic Science, 41, 1030-1040.

Sole, M., Porte, C. and Albaiges, J. (1995). Seasonal variation in the mixed function oxygenase system and antioxidant enzymes of the mussel Mytilus galloprovincialis. Environmental Toxicology and Chemistry, 14(1), 157-164.

Sole, M., Porte, C., Biosca, X, Mitchelmore, C.L., Chipman, J.K., Livingstone, D.R. and Albaiges, J. (1996). Effects of the "Aegean Sea" oil spill on biotransformation enzymes, oxidative stress and DNA-adducts in digestive gland of the mussel (Mytilus edulis L.). Comparative Biochemistry and Physiology, 113C(2), 257-265.

Sridevi, B., Reddy, K.V. and Reddy, S.L.N. (1998). Effect of trivalent and hexavalent chromium on antioxidant enzyme activities and lipid peroxidation in a freshwater field crab, Barytelphusa guerini. Bulletin of Environmental Contamination and Toxicology, 61, 384-390.

Stagg, R., Goksoyr, A., and Rodger (1992). Changes in branchial Na+, K+-ATPase, metallothionein and P450 1A1 in dab Limanda limanda in the German Bight: indicators of sediment contamination. Marine Ecology Progress Series, 91, 105-115.

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Sulaiman, N., George, S. and Burke, M.D. (1991). Assessment of sublethal pollutant impact on flounders in an industrialised estuary using hepatic biochemical indices. Marine Ecology Progress Series, 68, 207-212.

Suresh, K. and Mohandas, A. (1990). Hemolymph acid phosphatase activity pattern in copper-stressed bivalves. Journal of Invertebrate Pathology, 55, 118,125.

Svendsen, C. and Weeks, J.M. (1995). The use of a lysosomal assay for the rapid assessment of cellular stress from copper to the freshwater snail Viviparus contectus (Millet). Marine Pollution Bulletin, 31(1-3), 139-142.

Temara, A., Warnau, M., Dubois, Ph. and Langston, W.J. (1997). Quantification of metallothioneins in the common asteroid Asterias rubens (Echinodermata) exposed experimentally or naturally to cadmium. Aquatic Toxicology, 38, 17-34.

Thomas, P. and Wofford, H.W. (1993). Effects of cadmium Aroclor 1254 on lipid peroxidation, glutathione peroxidase activity, and selected antioxidants in Atlantic croaker tissues. Aquatic Toxicology, 27, 159-178.

Thompson, J.A.J., and Cosson, R.P. (1984). An improved electrochemical method for the quantification of metallothioneins in marine organisms. Marine Environmental Research, 11, 137-152.

Tom, M., Jakubov, E., Rinkevich, B. and Herut, B. (1999). Monitoring hepatic metallothionein mRNA levels in the fish Lithognathus mormyrus - evaluation of transition metal pollution in a Mediterranean coast. Marine Pollution Bulletin, 38(6), 503-508.

Torreblanca, A., Del-Ramo, J., Martinez, M., Diaz-Mayans, J., Pastor, A. (1996) Effect of 20-hydroxyecdysone administration on zinc, copper and metallothionein levels in Procambarus clarkii. Comp Biochem Physiol 113C (2) 201-204

van der Oost, R., Goksoyr, A., Celander, M., Heida, H. and Vermeulen, N.P.E. (1996). Biomonitoring of aquatic pollution with feral eel (Anguilla anguilla). II. Biomarkers: pollution-induced biochemical responses. Aquatic Toxicology, 36, 189-222.

Viarengo, A., Moore, M.N., Mancinelli, G., Mazzucotelli, A., Pipe, R.K. and Farrar, S.V. (1987). Metallothioneins and lysosomes in metal toxicity and accumulation in marine mussels: the effect of cadmium in the presence and absence of phenanthrene. Marine Biology, 94, 251-257.

Viarengo, A., Canesi, L., Pertica, M., Poli, G., Moore, M.N. and Orunesu, M. (1990). Heavy metal effects on lipid peroxidation in the tissues of Mytilus galloprovincialis LAM. Comparative Biochemistry and Physiology, 97C(1), 37-42.

Viarengo, A., Ponzano, E., Dondero, F. and Fabbri, R. (1997). A simple spectrophotometric method for metallothionein evaluation in marine organisms: an application to Mediterranean and Antarctic molluscs. Marine Environmental Research, 44(1), 69-84.

Viarengo, A., Burlando, B., Dondero, F., Marro, A. and Fabbri, R. (1999). Metallothionein as a tool in biomonitoring programmes. Biomarkers, 4(6), 455-466.

Widdows, J., Bakke,,T., Bayne, B.L., Donkin, P., Livingston, D.R., Lowe, D.M., Moore, M.N., Evans, S.V., Moore, S.L. Responses of Mytilus edulis on exposure to the water-accommodated fraction of North Sea oil. (1982) Mar Biol 67 (1) 15-31

Zielinski, S, and Portner, H.-O. (2000). Oxidative stress and antioxidative defence in cephalopods: a function of metabolic rate or age? Comparative Biochemistry and Physiology, 125 Part B, 147-160.

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3) Investigation into metallothionein induction in benthic organisms collected between Tyne/Tees and western Dogger Bank

3.1 : Introduction

The literature review had identified that of the techniques suggested by the JAMP cascade, the metal-specific biomarker relevant to UK marine waters was that of metallothionein (MT) induction. Accordingly, this biomarker was investigated in samples collected from the field.

The aim was to investigate the effectiveness of MT as a monitoring technique for biological effects of metals in marine waters. The hypothesis was that in areas with low concentrations of metals such as cadmium, copper and zinc, MT induction in organisms should be lower than in those collected from areas with higher concentrations of these metals. Accordingly, samples were collected between the western Dogger Bank and the Tyne/Tees region (Fig 3.1). The north east coast has elevated concentrations of cadmium in the sediments (Fig 1.2), which might be expected to induce MT in organisms living in the sediment. The Dogger Bank would normally be expected to be an uncontaminated, offshore area, but has shown anomalous results for metals such as cadmium eg relatively high concentrations of cadmium in dab livers (0.53 mg kg –1) (CEFAS, 1998). Therefore, the expectation was that MT induction ought to decrease with increasing distance from the coast, with the proviso that samples from the Bank itself might show anomalous results.

3.2 : Methods

As an experimental approach, MT induction was measured in benthos, since in comparison to fish, benthos should more fully reflect the surrounding sediments owing to their limited mobility. This leads to some difficulties since measurement techniques are most established for fish such as dab and flounder. Techniques for MT determination in edible mussels have also been developed, but these did not occur naturally across the study area. A further caveat was that limited, comparative data are currently available for benthic species. However, it was thought that the internal consistency of the sampling should allow any significant differences to be identified.

Trials were performed at CEFAS to investigate whether MT could be measured in species likely to be present in the sampling area, using a polarographic technique. These showed that species such as Asterias rubens, Echinocardium cordatum, and Pagurus spp. showed promise for MT induction measurements.

Opportunistic sampling was used to collect benthos and dab (Limanda limanda) by beam trawl in June 2000 on the RV Cirolana, between the western Dogger Bank (NMP station #285) to the north east English coast west of NMP station 245 (Fig 3.1). Summer sampling for MT measurement is not ideal because of seasonal variation, but was deemed acceptable here since the aim was to investigate relative rather than absolute differences. If significant differences in MT induction were observed, there would be evidence to support further optimisation of the methods.

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Fig 3.1(mapinfo plot not available in e-version)

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Biological samples were dissected for the appropriate organs on board ship, and were stored in cryovials at –80 ˚C until analysis in November-December 2001. Sediment samples were collected by Nioz corer in January 2001: it was unfortunate that they could not be collected at the same time as the trawl samples owing to bad weather conditions. Details of sample preparation are given in Annex 1.

Results reported here were obtained using a cadmium saturation technique, performed by Prof Stephen George’s group at the Institute of Aquaculture, University of Stirling. Data are reported as ng MT per mg homogenised protein extracted, as recommended under the BEQUALM programme. The analytical methods used are presented in Annex 2.

3.3 : Analytical quality control

Replicate metallothionein (3 replicates) and sample homogenate (3-4 replicates) determinations were carried out on all samples using the protocols listed in Annex 2. For the invertebrate hepatopancreas (digestive gland) samples, the homogenisation buffer was modified by addition of a digestive protease inhibitor (PMSF) to reduce protein degradation and by addition of 2-mercaptoethanol to reduce metallothionein oxidation. Assays of three concentrations of a flounder standard cytosol were performed for each batch of samples analysed as a quality control check. Replicate variances were <10%.

Total metal concentrations in sediments were subject to the same AQC procedures described more fully in section 5.3.

3.4 : Results

3.4.1 : Metals in sedimentsSampling stations are shown in fig 3.1 (yellow diamonds). Total metal concentrations in the sediments are shown in table 3.1, although data are not available for station 4. The values shown in table 3.1 show relatively low concentrations of cadmium and copper in sediments from stations 1-3. Zinc shows a little more variation being higher at station 2. Samples from station 5, close to the north east coast, show significantly higher concentrations in comparison with the other samples. If the sediment metal concentrations were the main control over the degree of MT induction, MT would be highest in those samples from station 2 and 5.

Table 3.1: Metal concentrations in sediments from MT sampling sites

Al %

Cd ng kg-1

Cr mg kg -1

Cu mg kg -1

Fe %

Mn mg kg -1

Pb mg kg -1

Zn mg kg -1

station1 1.30 <50 8 2 0.46 115 7 82 2.16 <50 14 3 0.83 157 14 383 2.24 <50 15 4 0.87 223 16 265 2.59 95 40 15 1.95 306 46 84

3.4.2 : Metallothionein results

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Samples were analysed for metallothionein and protein to allow quantification of MT as a proportion of the protein extracted. One of the criteria for analysis of biota to be analysed was their abundance across the sampling region. Of those analysed, Pagurus spp. were most common, 9-11 animals being present in trawls from each of the sampling stations. Asterias rubens, Echinocardium cordatum and dab each occurred at some of the stations, and it was expected that Pagurus spp. would provide a linkage between all the sampling sites. However, all the MT values in Pagurus samples were close to or below the detection limit of the Cd-satuaration assay, so additional work was undertaken to test the validity of the assay for this haemocyanin-containing animal. Haemocyanin is a copper-containing respiratory pigment found in the plasma and synthesised in the hepatopancreas. If copper is released from haemocyanin during the assay procedure (by proteolysis or heat treatment) and is oxidised, it displaces Cd and Zn from MT and therefore a Cd-saturation assay is unlikely to work unless the Cu1+ concentration is very low. Since low MT values were obtained with Pagurus, experiments were carried out where flounder liver supernatant samples were spiked with different concentrations of prawn (Macrobrachium) plasma. The results showed that the haemocyanin severely interfered with the estimation. Thus all results obtained for Pagurus were discarded as unreliable.

Table 3.2 shows the summarised analytical data for A. rubens, E cordatum and dab. Figure 3.2 shows mean concentrations and 1 standard error those for A. rubens and E cordatum, excluding the outliers (shown) which skew the mean value.

Table 3.2 : Summary statistics for MT ng mg-1 protein (including extreme values)

E. cordatum     A. rubens      station Mean se n min max Mean se n min max

1 113 15 8 31 158 11 3 3 6 172 186 43 5 22 269 8 * 1 * *3 87 26 11 21 315 * * * * *4 35 2 6 30 41 21 2 10 12 315 * * * * * 21 6 11 4 80

Dab (male)       Dab (female)    station Mean se n min max mean se n min max

1 345 46 8 131 559 813 310 4 302 17002 * * * * * * * * * *3 * * * * * * * * * *4 306 31 5 255 407 634 * 2 483 7865 * * * * * * * * * *

A. rubensThere were few differences in MT induction between the stations, although one sample collected from the western end of the transect at #5 showed a high value of 80 ng mg-1 (fig 3.2a).

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Fig 3.2

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E. cordatumUrchins were collected from stations 1-4, with the highest mean concentration of MT being found in samples from station 2. (A single sample from station showed a concentration of 315 ng mg-1.) Samples collected from station 1 on the Dogger showed elevated amounts of MT in comparison with those from #3-4 further west.

L. limandaDab were found at two of the sampling sites (1 and 4). Significant differences were not found between the two stations, but were apparent between the males and females at any given site.

3.5 : Discussion

Temara et al (1997) measured MT in A. rubens exposed to cadmium, and found background concentrations of 2.5-4.5 mg MT g –1 dry weight which increased to 5-5.6 mg MT g-1 dw in samples from a metal-contaminated fjord. While these units are not directly comparable with those used in the current work, the data suggest that A. rubens have the potential to be used for monitoring metal exposure. However, the data found in the current work do not show significant differences, and it may be that MT induction in this species is limited to areas where considerable ranges of sediment metal contamination exist.

Figure 3.2b shows that significant differences were obtained between the different stations for E. cordatum, and there is a suggestion that these broadly follow the sediment metal concentrations. Samples from station 1 showed slightly elevated MT in comparison with stations 3+4; while this might not be expected from the metals concentration, it is not inconsistent with data for dab and the history associated with the Dogger station. Any further work using this species would require clarification as to the reasons behind outliers eg the low value at station 2 and high value at station 3.

MT concentrations in dab provide useful comparative data since few such data are yet available for the benthos. Dab livers for MT analysis were collected by CEFAS as part of the NMMP in June 2001 (J Thain, pers. comm.), and these were also analysed by Prof George’s group at Stirling University. For male dab, MT concentrations in fish from the western Dogger were higher than or similar to those in fish collected further east. Concentrations were similar to, or lower than, samples collected elsewhere along the English east coast (Tees Bay to Outer Gabbard). In general, MT concentrations in female dab were higher than in males, possibly owing to seasonal variation associated with the reproductive cycle. The value of 813 ng mg-1 in the Dogger sample collected for the current work are higher than most of the NMMP samples, although this value is skewed by a very high value (1700 ng mg-1) in one fish. Repeat analysis showed that this value was real, so the sample may have been contaminated or the fish gravid.

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3.6 : Conclusions

MT induction has been investigated in Asterias rubens, Echinocardium cordatum, Pagurus spp. and Limanda limanda in samples collected from 5 stations between the western Dogger Bank and the Tyne/Tees coast.

Pagurus spp. are widely dispersed which is an advantage for in situ monitoring purposes. However, measurement of MT using the Cd-saturation assay used in the current study proved not to be possible, reducing the value of this species for biomonitoring.

The limited range of MT induction in A. rubens meant that few differences were found between stations of low-moderate metal concentrations in the sediments. Other workers (Temara et al, 1997) have examined MT in A. rubens in metal-contaminated areas and found significant differences in comparison to samples from less-contaminated areas.

MT induction in E. cordatum showed some association with sediment-metal concentrations, although further work is required to better understand MT dynamics in this echinoid.

MT in dab livers collected from the western Dogger showed similar or lower values in comparison with those collected under the NMMP from the English east coast.

3.7 : Reference

CEFAS (1998). Monitoring and surveillance of non-radioactive contaminants in the aquatic environment and activities regulating the disposal of wastes at sea, 1995 and 1996. Science Series, Aquatic Environment Monitoring Report, No. 51.

Temara A, Warnau M, Dubois Ph and Langston WJ (1997) “Quantification of metallothioneins in the common asteroid Asterias rubens” Aq Toxicol 38 17-34

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4) Marine phytoplankton – a mechanism for transferring dissolved cadmium to sediments – Literature review

4.1 : Summary

Over the last decade, various authors have produced results showing some elevated concentrations of cadmium in sediments and biota from the Dogger Bank region. Recent work (Whalley and Brown, unpub.) has suggested an association between particulate cadmium and phytoplankton in the water column. The current contract is seeking to identify whether this observed relationship provides a mechanism whereby dissolved cadmium may be taken up by primary producers, subsequently to fall to the sediments. This document reviews the evidence available in the literature for such a mechanism.

Growth of primary producers is primarily dependent on macronutrients such as phosphate and nitrate. However, trace nutrients such as some metals (eg iron, zinc) are also essential for population growth. Historically, cadmium (Cd) has had no known biological function and has been regarded as a toxic element. However, recent research has shown that marine phytoplankton may actively absorb cadmium under certain conditions, such as when zinc is limited in supply. It seems that marine algae may strip low concentrations of cadmium out of the water column and concentrate this metal into particulates of a bioavailable form.

Decaying plankton blooms provide a seasonal input of organic material to sediments. Cadmium associated with this material may thus be transferred to sediments, where it may be re-mineralised or ingested and absorbed by benthic feeders. Current research therefore indicates that there is a theoretical pathway for dissolved phase cadmium to be transferred into sediment and biota. Further practical work to be carried out on the current project seeks to investigate the pathway into sediments.

4.2 : Introduction

Until recently, there was no known biological function of cadmium (Cd), and it was regarded as a toxic element. However, its dissolved phase profile in open ocean waters showed similarities with those of nutrients such as phosphate (PO4), which caused some interest. Over the last decade, evidence has been mounting to show that cadmium may be being used by phytoplankton under some circumstances. As primary producers, phytoplankton form the basis of virtually all marine food webs. If Cadmium is being accumulated and used by plankton, there may be implications for higher trophic levels through subsequent bioaccumulation of the metal along food chains. In recent work undertaken for MAFF, C Whalley and J Brown (CEFAS) have found a correlation between Cadmium in suspended material and chlorophyll a, which is a marker for phytoplankton, in waters from the central North Sea (unpub. results). A further possible implication of cadmium acting as a nutrient has been highlighted by Wang and Dei (2001b), in which they note that metals may alter the response of phytoplankton growth rate to nutrient enrichment. Thus there are concerns revolving around both cadmium bioaccumulation and cadmium’s potential involvement in eutrophication.

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4.3 : Nutrients in seawater

Nutrients which are generally under consideration are the macronutrients such as nitrogen (as nitrate), phosphorus (as phosphate) and silicon (as silicate). However, also of relevance are trace elements such as iron (Fe), zinc (Zn) and copper (Cu). These elements, required in very low amounts, can be used in proteins or enzymatic processes. For example, Hutchins and Bruland (1994) investigated the biologically required trace elements zinc, iron and manganese (Mn) in diatom and flagellate prey, and their assimilation efficiencies (ie the amount of metal taken up) into crustacean grazers. They found that the metals behaved much like major nutrients during grazing. Where grazing exists, metal residence times in the water column are likely to increase (Twiss and Campbell, 1995) since the metals are being recycled.

4.4 : Plankton dynamics

For the purposes of this review, it is sufficient to know that at different stages in a plankton bloom, different types and species of plankton will dominate the biomass (or weight of living plankton material at a particular time), as diagrammatically depicted in Fig 1. In early stages of the bloom, there may be limited light availability or low temperatures, but excess nutrients. As time goes on, dissolved nutrients may become limiting and the phytoplankton community will be grazed by zooplankton, leading to a successional pattern similar to that depicted in Fig 1 for phytoplankton. Zooplankton comprise all free-living animals with mobility less than the physically-driven water transport (Båmstedt et al, 2000). In this work, we are interested in the fraction <5mm, which includes a range of invertebrates from groups such as copepods, foraminifera, flagellates, ciliates, larval and juvenile stages of benthos and fish.

Over large areas of the European Continental Shelf during summer the thermocline in thermally stratified waters marks a distinct interface between nutrient-rich, denser, colder waters below and nutrient-depleted, less dense upper waters. In these areas, dissolved nutrients in surface waters tend to be depleted early in the growing season (Barber and Smith, 1981). Phytoplankton continue to grow at the thermocline, where there is sufficient light and an upward flux of nutrients. In surface waters, low levels of primary production continue utilising nutrients recycled within the ecosystem.

The plankton bloom is generally very efficient at recycling nutrients within the biomass. However, losses of nutrients can occur through dead organisms or fecal pellets dropping to the sediment, or when fish (for example) feed on the plankton and then move away. Sedimented particles may be eaten by detrital or sediment feeders, or remineralised. Recycling of nutrients is of relevance to this study since this may provide a mechanism whereby dissolved phase cadmium can be incorporated into the “food chain”.

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4.5 : Toxicity of cadmium to phytoplankton

Toxic effects have an inhibitory effect on productivity of the plankton bloom, which results in a reduced biomass. This in turn reduces the amount of metal accumulated (Murray and Maskell, 1986). Much work has been directed towards investigations of toxic effects of cadmium to aquatic organisms (eg Fisher et al , 1984; Kuiper, 1981; Payne and Price, 1999). Metal toxicity to phytoplankton may occur by enzyme inactivation, transport interference and major nutrient mimicry or modification (Price and Morel, 1994).

Brand et al (1986) examined cadmium toxicity in phytoplankton, finding that toxicity increased according to diatoms < coccolithophoroids ≈ dinoflagellates < cyanobacteria. Price and Morel (1994) observed that cadmium interfered with iron transport and metabolism. Nutrient mimicry is demonstrated by Cd2+, which has a similar charge density to that of Ca2+ so it can substitute into the crystal lattice of CaCO3 mineral formed by foraminifera (Frew and Hunter, 1992). Trace metal interactions can become complicated. For instance, Sunda and Huntsman (1996, 1998) reported antagonistic relationships in a diatom between manganese, copper, cadmium and zinc, with, for example, effects being caused by zinc or cadmium blocking uptake of manganese in Mn-uptake sites, transport of eg cadmium into the cell via the Mn-uptake system, and Mn-uptake inhibition as a response to detoxification mechanisms.

4.6 : Cadmium as a nutrient

For some years, there have been reports that dissolved cadmium profiles in marine waters are similar to those of macronutrients like nitrate and phosphate (Bruland, 1980; Frew and Hunter, 1992). Lee and Morel (1995) considered cadmium to be an algal nutrient. Acceptance of this behaviour has been marked recently in the OSPAR QSR 2000 report with the statement that “Cadmium behaves like a nutrient in seawater” (p47, OSPAR Commission, 2000).

The relationship between cadmium and PO4 is not always observed. Hall et al (1999) found that the cycling of cadmium and PO4 were de-coupled in a Scottish sea loch. They attributed this partly to different uptake and re-generation rates of cadmium and PO4, and partly to the relatively high concentrations of dissolved cadmium in shelf seas and coastal waters which means that uptake of cadmium by plankton is too small to be seen. While recognising an association between the behaviour of PO4 and cadmium, Löscher et al (1998) did not see the expected cadmium depletion in surface waters of the Southern Ocean. They attributed this observation to upwelling of Upper Circumpolar Deep Water, which has relatively high cadmium concentrations, and to low biological productivity. Frew and Hunter (1992) noted that the global dissolved Cd:PO4 ratio was lower in the Cd-depleted waters of the Southern Ocean. Schneider and Pohl (1996) examined dissolved cadmium at coastal station in the western Baltic Sea. They tentatively suggested a de-coupling of dissolved cadmium from nutrients during the spring bloom, because of a 2-3 month time shift in the dissolved nutrient and cadmium profiles.

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Although there are some studies where the Cd-PO4 association has not been observed, these tended to be in coastal regions where there were relatively high metal concentrations (including zinc, for which it can substitute). These could make it difficult to see small changes in dissolved concentrations. The Southern Ocean study showed another possible “interference”, where a water input disguises any observations of small-scale concentration differences (Schneider and Pohl, 1996). Otherwise, it seems that the dissolved phase association between cadmium and PO4 is an accepted phenomenon.

4.7 : Cadmium uptake by phytoplankton

The rate of metal uptake from the surrounding water will depend upon the metal, the species and the ambient conditions. Fisher et al (1984) looked at cadmium, zinc, silver and mercury uptake by a diatom, a chlorophyte, a coccolithophore and a cyanophyte. They found that metal uptake in dead cells was initially the same as that in living cells, indicating that adsorption is the first association between the cell and the metal. This passive adsorption is followed by absorption of the sorbed metal through the cell walls against a concentration gradient (Murray and Maskell, 1986). Cadmium bound to colloids was found to enhance the initial surface sorption of cadmium to algal cells (Wang and Guo, 2000). Colloids are common in fresh and coastal waters, but are generally of negligible concentration offshore. Murray and Maskell (1986) summarised the then current knowledge for modelling by assuming that all phytoplankton are in equilibrium with the dissolved metals in surrounding seawater. Such an approach does not consider instances of uptake against a concentration gradient.

Dinoflagellates may have a greater ability to accumulate metals by adsorption than diatoms as a result of their flagella (ie greater surface area) (Murray and Maskell, 1986). These authors proposed that for modelling purposes, the greatest refinement necessary for species differentiation was to distinguish between diatoms and dinoflagellates. Large differences in adsorption capacity could account for seasonal variation in spring and summer dinoflagellate blooms.

If all metals were simply sorbed through the cell walls, uptake rates would be similar for all metals. However, in the case of cadmium, there appears to be a more active uptake mechanism. This may be a result of it acting as an analogue for the essential element zinc, as a result of its similar electron structure. Luoma et al (1998) found that a relatively large proportion of these two metals, relative to other metals, is taken up into a labile, intracellular fraction of phytoplankton analogous to some nutrients. The same authors noted that in an estuary, the biological impact on cadmium and zinc on cycling was marked, while any effects on copper were negligible.

Price and Morel (1994) suggested that the metal content of phytoplankton may be more a reflection of intracellular metal metabolism than that of the roles metals play in biochemical pathways. This was possibly a result of the slow kinetics of metal incorporation into enzymes and proteins, forcing the plankton to store metals within the cells.

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Suspicions that cadmium might substitute for zinc in Zn-limited phytoplankton were based on the behaviour of metalloenzymes in vitro and in vivo (Price and Morel, 1990). Although zinc concentrations in sea water are much higher than those of cadmium, zinc tends to be more extensively organically complexed than cadmium, and thus can be unavailable for biological uptake.

Empirical evidence that cadmium could substitute for zinc was first presented in 1990 (Price and Morel, 1990). These authors showed that cadmium could stimulate the growth of the diatom Thalassiosira weissflogii by substituting for zinc in certain macromolecules in a laboratory study. Cobalt showed similar, but less effective, zinc substitution abilities. Lee and Morel (1995) examined the Cd/Zn substitution effects in 9 types of phytoplankton, including both oceanic and coastal species. They found that cadmium could enhance the growth of a variety of species, under environmentally relevant inorganic cadmium and zinc concentrations. It should be noted that these laboratory studies are difficult to perform, since ensuring there is no zinc contamination and keeping phytoplankton alive are both specialised tasks.

In more detailed work, Lee et al (1995) investigated the range of concentrations at which Cd/Zn substitution occurred. If there was no added zinc (estimated concentration <0.1pM = ~1.5 fg l-1), T. weissfloggii was unable to grow and 4.6 pM added cadmium (~41 fg l-1) was lethal. At least 2 pM Zn was required for growth, and under these conditions, 4.6 pM Cd enhanced growth of zinc-limited cells. However, at cadmium concentrations greater than 23 pM, toxic effects negated the beneficial effects. At a higher zinc concentration (16 pM) the effective range for cadmium additions was smaller, although they also did not see the toxic effects observed at the lower zinc concentration.

At high cadmium concentrations, there was a shift in the cadmium location in T. weissflogii from membrane-bound material to the cytoplasm, possibly as a result of production of a Cd-phytochelatin complex as a detoxification mechanism (Lee et al, 1995). ~40% of cadmium in the cell was bound to the membrane while the rest was fractionated into the fecal material. Of the latter, some could sink to the sediment and be re-mineralised.

Lee et al (1995) reported that cadmium enhances the activity of the zinc metalloenzyme carbonic anhydrase (CA). A mechanism for the utilisation of cadmium by phytoplankton has recently been reported by Cullen et al (1999). They suggested that a cadmium – carbonic-anhydrase is formed. This Cd-CA is inversely related to the pCO2 and the zinc concentration, so more dissolved zinc reduces the cadmium uptake.

Wang and Dei (2001a) investigated the impact of adding nutrients to the diatom Thalassiosira pseudonana on its metal uptake rate. Addition of nitrogen significantly increased the amount of cadmium taken up (but not that of selenium or zinc), while silicate reduced the selenium uptake. Phosphate had no significant effect on metal uptake. These are interesting findings since they suggest that uptake of cadmium may be increased under high-nutrient conditions. This may then have subsequent impact upon potential toxicity and metal transfer into the aquatic food chain. In further work (Wang and Dei, 2001b), the authors investigated four species of marine phytoplankton (diatom, green alga, dinoflagellate and prasinophyte) and the influence of nutrients on

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metal uptake over a 5 hour period. They found that addition of nitrogen at 176.4 µM (12.6 µg), caused the dry weight concentration factor for cadmium to increase 2-15 times, with 1-4 times increase for selenium and 1-5 times for zinc. Additions of phosphate or silicate showed less pronounced effects. They also found a correlation between cell growth rate and metal uptake, although no causal link was identified.

4.8 : Trophic transfer and sedimentation

One implication of phytoplankton actively absorbing cadmium is that low dissolved metal concentrations may be scavenged and concentrated into particles. These particles may then be ingested by other organisms in the water column, or sedimented out as detrital material. On the sea bottom, organic particles may either be scavenged or re-mineralised. The following section examines the literature for evidence of trophic transfer and sedimentation of cadmium incorporated into phytoplankton.

Zauke et al (1996) reported the range of cadmium concentrations in copepods as being from 0.13 mg/kg (dry weight) in fish larvae to 51 mg/kg in hyperiid amphipods. Enrichment factors of cadmium in the copepod Calanus finmarchius collected in the northern North Sea and north eastern Atlantic compared to the dissolved concentration were ~ 2, similar to that in the SPM generally (Haarich et al, 1993). Whalley et al (1997) found cadmium concentrations in plankton (200-1000 µm) from the Dogger Bank region to be between 0.5-2.5 µg g-1 (dry weight). Changes in metal accumulation (Zn, Cu and Cd) were related to plankton activity in a study performed by Wolter et al (1984). They found that accumulation of metal by phytoplankton, expressed as an enrichment factor, was greatest at the lowest ambient metal concentration, with a plateau being reached at higher concentrations.

Luoma et al (1998) studied metal uptake during a bloom and found that dissolved cadmium was reduced to 50% of its pre-bloom concentration in an estuary. They found that the mass of cadmium taken up by the phytoplankton was similar to the mass of cadmium removed from solution if particle settling were considered. In the polar region, Bargagli et al (1996) found that cadmium concentrations in the sediments were similar to those in other background areas. However, the values in surface waters, phyto- and zoo-plankton were similar to those found in areas of enhanced upwelling. They suggested that the elevated cadmium concentrations were a result of rapid cadmium regeneration, and found that the metal was accumulating in molluscs. Kuiper (1981) added cadmium to experimental “ecosystem bags” and found that little cadmium (5-9%) accumulated in the sediment over the 3-month experiment. He attributed this partly to the relatively low partitioning coefficient of cadmium to solids in seawater, and partly to the regeneration of nutrients.

Seasonal variation in metal and carbon, nitrogen and phosphorus concentrations was measured in material caught in sediment traps moored 20m above the sea bottom in the Baltic Sea (Leivuori and Vallius, 1998). In a 6-month study, metal accumulation in the traps was lowest during the spring bloom but subsequently started rising, with the highest metal concentrations being found in material from the autumn bloom. Hall et al (1999) examined cadmium in a Scottish sea loch in relation to phytoplankton blooms. Measurements of particulate, leachable cadmium in sediment traps showed a lag of ~2 weeks in apparent uptake and regeneration relative to those

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of phosphate. Maximum enrichment (>8 times) compared to the winter cadmium concentrations occurred 1-2 weeks after the bloom had crashed.

Metal introduction to sediments can exhibit a temporal relationship with periods of primary productivity (Schlekat et al, 2000). Shaw et al (1994) found productivity-dependent scavenging/uptake of uranium in surface waters. They noted that this mechanism provides a rapid transport route for dissolved metals in the water column to be delivered to the sediments. Schneider and Pohl (1996) suggested that precipitation of fecal pellets from zooplankton were more important in transferring cadmium to sediments than phytoplankton sedimentation.

Schoemann et al (1998) studied dissolved iron and manganese in shallow, coastal waters in the North Sea. They found that the dissolved concentrations increased at the end of the spring which they attributed to sedimentation of aggregated phytoplankton-derived material. The increase was observed because settling of fresh organic material would have stimulated benthic microbial activity, which would then alter the redox conditions in the sediment. Reducing conditions lead to the dissolution of precipitated Fe and Mn and hence their appearance in the dissolved form in the overlying water. The authors concluded that the biogeochemical cycles of iron and manganese are driven by the eutrophication-dependent magnitude of phytoplankton blooms, the heterotrophic activity that follows the spring bloom and the quality of phytoplankton-derived material that reaches the sediment.

In a study in freshwater lake sediments, Warren et al (1998) found that generally, benthic uptake of cadmium was more closely associated with the concentration in porewaters rather than in the sediment. However, sediment feeding organisms (as opposed to those merely inhabiting the sediment) did obtain substantial amounts of cadmium from the sediment.

Lee and Luoma (1998) fed bivalves (clams) on suspended material collected either during a phytoplankton bloom (high organic content) or during a low phytoplankton period when the suspended material was dominated by re-suspended sediment (lower organic content). They spiked the material with radioisotopes and looked at the absorption efficiency of contaminants. Greater absorption of cadmium and chromium occurred with greater proportion of organic material (microalgae) present. The authors attributed this behaviour to the higher bioavailability of metals in cytosolic form compared to those bound to sediment. They estimated that phytoplankton blooms may result in up to a twofold increase in cadmium and zinc bioaccumulation from food suspensions, although the actual extent is dependent upon metal and species.

Cadmium adsorbed on dead algae represented a highly available source of cadmium for benthic invertebrates (such as Abra alba, and to a lesser extent, Amphiura filiformis and Mytilus edulis) feeding on detritus (Schaanning et al, 1996). In contrast, Schlekat et al (2000) found that the estuarine amphipod Leptocheirus plumulosus had a low assimilation efficiency of cadmium in phytoplankton. This organism took in most cadmium from bacterial exopolymeric coatings, ie labile, polymeric organic carbon sediment coatings. Lee et al (1995) estimated that 40% of phytoplankton-cadmium is available to be assimilated into copepods grazing on the algae.

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Ni et al (2000) showed that fish preying on zooplankton were accumulating metals such as cadmium, chromium and zinc, with trophic transfer of contaminants being the outcome of their study. The assimilation efficiency of cadmium to glassy (a fish) was 14-23 %, and to mudskipper was 10-26 %. Trophic transfer of cadmium from plankton to barnacles was also identified by Wang et al (1999).

Cadmium is not always associated with the phytoplankton. Miramand et al (1993) did not find a relationship between cadmium or vanadium and plankton blooms in the Bay of Seine. Showell and Gaskin (1992) examined cadmium in seston, in organic detrital material (as particulate organic carbon), suspended clays (as aluminium) and phytoplankton (measured by chlorophyll a) components. They found that cadmium was mostly associated with organic detrital material (although this could be decayed phytoplanktonic material).

Overall, the literature suggests that cadmium uptake by phytoplankton does provide a route for this metal to enter biota from initial low dissolved concentrations. The metal can be scavenged by phytoplankton from the water column, concentrated into the organisms, and then be sedimented out as detrital material. Trophic transfer into benthic organisms can occur where the organic material has fallen to the sediment. This may occur to a significant extent when a bloom crashes. Less evidence is available for trophic transfer of cadmium from phytoplankton to fish, although it does appear that this can occur.

4.9 : Metal flux by plankton blooms

There was little discussion of this topic in the literature, which is complicated by the use of the word “transport” in relation to plankton in an intracellular sense. As an example of the work possible, Murray and Maskell (1986) estimated that in Liverpool Bay, the amount of dissolved copper amounted to 900t, copper on suspended solids was 300t and that in phytoplankton cells was 10t. While the absolute amount of copper bound up in the plankton was relatively small, this portion of the metal was being continuously recycled, and was thus part of the smaller, “active pool” of metal. Similar characteristics may be attributed to other metals.

4.10 : Production in the Dogger Bank region

The deeper waters of the central North Sea (>40m) are thermally stratified (warm water overlying cool) from May to October. In the vicinity of the comparatively shallow (< 30 m) Dogger Bank the stratification is intermittent, being largely dependent on the degree of wind mixing. The interface between the warm and cool waters lies at typically 30 – 40 m and acts to isolate nutrient-rich winter water below a nutrient-depleted upper layer. Additionally, a series of bottom fronts surround the Dogger Bank, forming a transition from the deep stratified waters to the warmer waters above the Bank. The position of these fronts is predictable and determined by a balance between the degree of tidally induced stirring of the water column and solar heating providing buoyancy to the surface layers (e.g. Brown et al, 1999; Brown et al, 2001). In the deeper waters where tides are too weak to mix down warm surface

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layers the water column stratifies. At the interface between the warm and cold water there is sufficient light and nutrient supply to permit phytoplankton growth (e.g. Richardson et al, 1998).

In addition, the bottom fronts that separate the deeper stratified regions of the central North Sea from shallower mixed waters generate a comparatively narrow (< 20 km) ‘jet-like’ circulation with flows exceeding 15 cm s-1 (Brown et al, 1999; Brown et al, 2001). The flows persist throughout the period of stratification, extending continuously for ~ 500 km from the Firth of Forth to Flamborough Head before passing offshore to the Dogger Bank.

Neilsen et al (1993) investigated plankton in the Dogger Bank area in early summer 1990. They found that on and to the south of the Bank, waters were nutrient depleted. In these areas, approximately 15% of phytoplankton production was channelled directly into zooplankton (copepods). In the stratified waters to the north of the Bank, ~30% of the primary production was ingested by copepods. With increasing distance north, production was based on a succession towards a community production using re-generated nutrients.

Production in the shallow waters of the Dogger Bank continues throughout the year. In these waters, there is high bacterial biomass and associated production just above the bottom. This may result from low retention efficiency of particles <1 µm by benthic suspension feeders and their release of dissolved organic nitrogen (Neilsen et al 1993). These authors also suggested that new production along the edge of the Bank may arise from the intrusion of nutrient-rich water across the pycnocline. Recent work by Cannaby et al (2000) directly measured the movement of water between the dense, nutrient-rich cold pool on the northern flank of the Bank into the warmer, upper layers. Velocities were extremely weak and apparently insufficient to carry nutrients directly along the thermocline. However, it is likely that there is sufficient diffusion of nutrient across the thermocline to sustain production.

Trimmer et al (1999) examined the spring bloom in a similarly summer-stratified region of the western Irish Sea. They assessed the degree of coupling between water column production during the bloom and benthic processes by building a carbon budget. Their estimate was that the spring bloom comprised 76% new production and 24% regenerated production. Approximately 40% of the spring production (ie ~50% of the new production) was inputted to the benthos, as estimated by carbon and oxygen measurements. Their results suggested that there is considerable inter-annual variability in the importance of the spring bloom to secondary benthic production. When the spring bloom was of low significance, they suggest that benthic production must be supported by phytoplankton production which occurs after the bloom or by detrital organic carbon advected into the area after breakdown of stratification. Similar conditions may exist at the Dogger.

In their study of the plankton community in the Dogger Bank region, Richardson et al (1998) observed both vertical and horizontal heterogeneity, producing a “patchy distribution”. There was a subsurface chlorophyll peak as previously observed by these workers (Neilsen et al, 1993) and by others (J. Brown, pers. comm.), which occurs where nutrient-rich bottom waters meet the photic zone (which is otherwise nutrient-depleted). Richardson et al (1998) predicted that in the areas of subsurface

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chlorophyll peaks, there will be efficient transfer of energy from phytoplankton to higher trophic levels, together with accumulations of pelagic feeders (fish and larvae).

4.11 : Cadmium in sediments from the Dogger Bank region

Cadmium concentrations in sediments at the Dogger have been studied in detail by DFR/CEFAS (Rowlatt and Lovell, 1994; Whalley et al, 1997). In the earlier work, a few elevated cadmium concentrations were found in sediments from the Dogger Bank. This prompted further work (funded by DoE) where 213 sediment samples from the region of the Bank were collected and analysed for metals. Again, there were some elevated cadmium concentrations, at up to 0.24 µg g-1 in sands. Such concentrations were unexpected because sands do not normally accumulate contaminants. Analytical reproducibility was good; however, sediment sampling reproducibility was more variable. In samples collected offshore from replicate grabs, cadmium concentrations in those west of the Netherlands were in the range 0.02 – 0.05 µg g-1 (n= 6, sd = 38%), while those collected from the western Dogger Bank showed greater variability at <0.02 – 0.18 µg g-1 (n = 6, sd = 100%). This variability might be explained by a variable input of cadmium from settled clumps of decaying plankton, and is the subject of further work under the current contract.

4.12 : Cadmium in benthos from the Dogger Bank region

Dab livers collected from fish from the Dogger Bank have a history of elevated cadmium concentrations (CEFAS, 1998). This observation has prompted several investigations into concentrations of cadmium in biota from the Bank (Stebbing and Dethlefsen, 1992; Cofino et al, 1992; Whalley et al, 1997). Concentrations of cadmium in Aphrodite aculeata were < 1 µg g-1 (dry weight) in samples collected from on and south of the Bank, but were 1-4 µg g-1 in those collected off the northern side of the Bank (Whalley et al, 1997). A similar pattern was observed for concentrations in Asterias rubens, but not in Pagurus spp. Langston et al (1999) looked at accumulation of metal contaminants in benthos incubated in sediment cores collected from the Dogger Bank. They found that although absolute concentrations of cadmium were low, the metal was accumulated by bivalves, a gastropod and a polychaete. Bioaccumulation was attributed to the cadmium being in a labile form.

4.13 : Conclusions

Results from studies into cadmium in sediments and biota from the Dogger Bank region have shown some elevated concentrations in various studies over the last decade. Work by Whalley and Brown in this region, carried out under contract to MAFF, has shown an association between cadmium and chlorophyll a in suspended material, suggesting a relationship with phytoplankton. The current contract is seeking to identify whether this observed relationship provides a mechanism whereby dissolved cadmium may be taken up by primary producers, subsequently to fall to the sediments. This document reviews the evidence available in the literature for such a mechanism.

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Cadmium does appear to be actively taken up by marine phytoplankton in certain circumstances, such as when zinc is limiting. This behaviour has the effect of concentrating up low dissolved concentrations of the metal into small particles. These particles may then either be recycled within the bloom, or exported either to the sea bed in detrital material or out of the system in pelagic feeders. Trophic transfer of cadmium is possible, providing a mechanism for dissolved cadmium to enter the “food chain”.

Plankton blooms are “patchy” in space, as are any settled clumps of material from decaying blooms, which could account for the spatial variability seen in cadmium concentrations in sediments at the Dogger Bank.

The ability of phytoplankton to strip cadmium out of the water column has implications for sedimentation and trophic transfer of this metal. Any effects may be of less significance in coastal waters where other trace nutrients are likely to be available and concentrations are generally higher. However, blooms in offshore regions may represent an as yet unquantified biogeochemical input into the sediments, with potential for trophic transfer into benthic organisms.

4.14 : References

Båmstedt U, Gifford DJ, Irigoien X, Atkinson A and Roman M. (2000) “Feeding” IN ICES Zooplankton Methodology Manual ed RP Harris, PH Wiebe, J Lenz, HR Skjoldal and M Huntley Academic Press, London ISBN 0 12 327645 4

Barber RT and Smith RL. (1981) “Coastal upwelling ecosystems” IN Analysis of Marine Ecosystems ed. AR Longhurst. Academic Press, London. ISBN 0 12 455560 8

Bargagli R, Nelli L, Ancora S and Focardi S. (1996) “Elevated cadmium accumulation in marine organisms from Terra Nova Bay (Antarctica)” Polar Biol 16 (7) 513-520

Brand LE, Sunda WG and Guillard RR. (1986) “Reduction of marine phytoplankton reproduction rates by copper and cadmium” J Exp Mar Biol Ecol 96 225-250

Brown J, Hill AE, Fernand L and Horsburgh KJ. (1999) “Observations of a Seasonal Jet-like Circulation at the Central North Sea Cold Pool Margin” Est Coast Shelf Sci 48 (3) 343-355

Brown J, Fernand L, Horsburgh KJ, Hill AE. and Read JW. (2001). “Paralytic shellfish poisoning on the east coast of the UK in relation to seasonal density-driven circulation” J Plankton Res 23 105-116

Bruland K. (1980) “Oceanographic distributions of cadmium, zinc, nickel and copper in the north Pacific” Earth Planet Sci Lett 47 176-198

Cannaby H, Fernand L, Horsburgh KJ, Brown J, Tinton EJ, Read JW and Hill AE. (2000) “A tracer experiment to illustrate the role of frontal circulation in the supply of nutrients to thin phytoplankton layers in shelf sea thermoclines” UK Marine Science 2000. University of East Anglia.

CEFAS. (1998) Monitoring and surveillance of non-radioactive contaminants in the aquatic environment 1995 and 1996. CEFAS, Lowestoft. AEMR 51 ISSN 0142 2499

Cofino, WP, Smedes F, de Jong SA, Abarnou, A, Boon JP, Oostingh I, Davies IM, Klungsoyr J, Wilhelmsen S, Law RJ, Whinnett JA, Schmidt D and Wilson S. (1992) "The chemistry programme" Mar Ecol Prog Ser 91 (1-3) 47-56.

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Cullen JT, Lane TW, Morel FM and Sherrell RM. (1999) “Modulation of cadmium uptake in phytoplankton by seawater CO2 concentration” Nature 402 165-167

Fisher NS, Bohé M and Teyssié J-L. (1984) “Accumulation and toxicity of Cd, Zn, Ag and Hg in four marine phytoplankters” Mar Ecol Prog Ser 18 201-213

Frew R and Hunter D. (1992) “Influence of Southern Ocean waters on the cadmium-phosphate properties of the global ocean” Nature 360 144-146

Haarich M, Kienz W, Krause M, Zauke G-P and Schmidt D. (1993) “Heavy metal distribution compartments of the northern North Sea and adjacent areas” Dt Hydrogr Z 45 313-336

Hall IR, Hydes DJ, Statham PJ and Overnell J. (1999) “Seasonal variations in the cycling of aluminium, cadmium and manganese in a Scottish sea loch” Cont Shelf Res 19 1783-1808

Hutchins DA and Bruland KW. (1994) “Grazer-mediated regeneration and assimilation of Fe, Zn and Mn from planktonic prey” Mar Ecol Prog Ser 110 (2-3) 259-269

Kuiper J. (1981) “Fate and effects of cadmium in marine plankton communities in experimental enclosures” Mar Ecol Prog Ser 6 (2) 161-174

Langston WJ, Burt GR and Pope ND. (1999) “Bioavailability of metals in sediments of the Dogger Bank (central North Sea): a mesocosm study” Est Coast Shelf Sci 48 519-540

Lee B-G and Luoma SN. (1998) “Influence of microalgal biomass on absorption efficiency of Cd, Cr and Zn by two bivalves from San Francisco Bay” Limnol Oceanogr 43 (7) 1455-1466

Lee J and Morel FM. (1995) “Replacement of zinc by cadmium in marine phytoplankton” Mar Ecol Prog Ser 127 305-309

Lee J, Roberts S and Morel FM. (1995) “Cadmium: a nutrient for the marine diatom Thalassiosira weissflogii” Limnol Oceanogr 40 (6) 1056-1063

Leivuori M and Vallius H. (1998) “A case study of seasonal variation in the chemical composition of accumulating suspended sediments in the Gulf of Finland” Chemosphere 36 (3) 503-521

Lewis WM. (1979) Zooplankton Community Analysis. Springer-Verlag, New York ISBN 0 387 90434 4

Löscher BM, de Jong JT and de Baar HJ. (1998) “The distribution and preferential biological uptake of cadmium at 6ºW in the Southern Ocean” Mar Chem 62 259-286

Luoma SN, van Geen A, Lee B-G and Cloern JE. (1998) “Metal uptake by phytoplankton during a bloom in South san Francisco Bay” Limnol Oceanogr 43 (5) 1007-1016

Miramand P, Bentley D, Guary J-C and Brylinski J-M. (1993) “Role of plankton on the biogeochemical cycle of cadmium and vanadium in the eastern area of the bay of Seine” Oceanol Acta 16 (5-6) 625-632

Murray GE and Maskell JM. (1986) Heavy metal transport by phytoplankton in tidal waters. Hydraulics Research, Wallingford. Report no. SR87

Nielsen TG, Løkkegaard B, Richardson K, Pedersen FB and Hansen L. (1993) “Structure of plankton communities in the Dogger Bank area (North Sea) during a stratified situation” Mar Ecol Prog Ser 95 115-131

Ni I-H, Wang W-X and Tam YK. (2000) “Transfer of Cd, Cr and Zn from zooplankton prey to mudskipper and glassy fishes” Mar Ecol Prog Ser 194 203-210

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OSPAR Commission. (2000) Quality Status Report 2000. Region V – Wider Atlantic. OSPAR Commission, London. 0 946956 51 0

Payne CD and Price NM. (1999) “Effects of cadmium toxicity on growth and elemental composition of marine phytoplankton” J Phycol 35 (2) 293-302

Price NM and Morel FM. (1990) “Cadmium and cobalt substitution for zinc in a marine diatom” Nature 344 658-660

Price NM and Morel FM. (1994) “Trace metal nutrition and toxicity in phytoplankton” IN Algae and Water Pollution ed LC Rai, JP Gaur and CJ Soeder 79-97

Richardson K, Nielson TG, Pedersen FB, Heilmann JP, Løkkegaard B and Kaas H. (1998) “Spatial heterogeneity in the structure of the planktonic food web in the North Sea” Mar Ecol Prog Ser 168 197-211

Rowlatt S and Lovell D. (1994) Survey of contaminants in coastal sediments. MAFF, Burnham-on-Crouch. DoE research contract PECD 7/7/358

Schaanning MT, Hylland K, Eriksen D, Bergan TD, Gunnarson JS and Skei J. (1996) “Interactions between eutrophication and contaminants II. Mobilisation and bioaccumulation of Hg and Cd from marine sediments” Mar Poll Bull 33 (1-6) 71-79

Schlekat CE, Decho AW and Chandler GT. (2000) “Bioavailability of particle-associated silver, cadmium and zinc to the estuarine amphipod Leptocheirus plumulosus through dietary ingestion” Limnol Oceanogr 45 (1) 11-21

Schneider B and Pohl C. (1996) “Time series for dissolved cadmium at a coastal station in the western Baltic sea” J Mar Syst 9 (3-4) 159-170

Schoemann V, de Baar HJ, de Jong JT and Lancelot C. (1998) “Effects of phytoplankton blooms on the cycling of manganese and iron in coastal waters” Limnol Oceanogr 43 (7) 1427-1441

Shaw TJ, Sholkowitz ER and Klinkhammer G. (1994) “Redox dynamics in Chesapeake Bay” Geochim Cosmochim Acta 58 (14) 2985-2995

Showell MA and Gaskin DE. (1992) “Partitioning of cadmium and lead within seston of coastal marine waters of the western Bay of Fundy, Canada” Arch Environ Contam Toxicol 22 325-333

Stebbing AR and Dethlefson V. (1992) "Introduction to the Bremerhaven workshop on biological effects of contaminants" Mar Ecol Prog Ser 91 (1-3) 1-8.

Sunda WG and Huntsman SA. (1996) “Antagonisms between cadmium and zinc toxicity and manganese limitation in a coastal diatom” Limnol Oceanogr 41 (3) 373-387

Sunda WG and Huntsman SA. (1998) “Interactions among Cu2+, Zn2+, and Mn2+ in controlling cellular Mn, Zn and growth rate in the coastal alga Chlamydomonas” Limnol Oceanogr 43 (6) 1055-1064

Trimmer M, Gowen RJ, Stewart BM and Nedwell DB. (1999) “The spring bloom and its impact on benthic mineralisation rates in western Irish Sea sediments” Mar Ecol Prog Ser 185 37-46

Twiss MR and Campbell PG. (1995) “Regeneration of trace metals from picoplankton by nanoflagellate grazing” Limnol Oceanogr 40 (8) 1418-1429

Wang W-X and Dei RC. (2001a) “Metal uptake in a coastal diatom influenced by major nutrients (N, P and Si)” Wat Res 35 (1) 315-321

Wang W-X and Dei RC. (2001b) “Effects of major nutrient additions on metal uptake in phytoplankton” Env Poll 111 233-240

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Wang W-X and Guo L. (2000) “Bioavailability of colloid-bound Cd, Cr and Zn to marine plankton” Mar Ecol Prog Ser 202 41-49

Wang W-X, Qiu J-W and Qian P-Y. (1999) “The trophic transfer of Cd, Cr and Se in the barnacle Balanus amphitrite from planktonic food” Mar Ecol Prog Ser 187 191-201

Warren LA, Tessier A and Hare L. (1998) “Modelling cadmium accumulation by benthic invertebrates in situ.” Limnol Oceanogr 43 (7) 1442-1454

Whalley C. (1995) Estimating binding strength and chemical phase of metals adsorbed to sediment components. PhD thesis. University of East Anglia, UK.

Whalley C, Rowlatt S, Jones L, Bennett M and Campbell S. (1997) Metals in sediments and benthos from the Dogger Bank, North Sea. CEFAS Burnham-on-Crouch, DoE contract CW0 301

Wolter K, Rabsch U, Krischker P and Davies A. (1984) “Influence of low concentrations of cadmium, copper and zinc on phytoplankton of natural water samples” Mar Ecol Prog Ser 19 (1-2) 167-173

Zauke G, Krause M and Weber A. (1996) “Trace metals in mesozooplankton in the North Sea” Int Rev Gesamt Hydrobiol 81 (1) 141-160

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5) Investigations into cadmium and chlorophyll a concentrations in sediments from the Dogger Bank region

5.1 : Introduction

The practical work was primarily aimed at examining whether observed particulate cadmium-chlorophyll a associations in the water column could be directly linked with those in the surface sediments of the Dogger Bank. If an association could be found, it would provide evidence towards a natural process inputting cadmium into the sediments in this region of relatively high and persistent primary production.

5.2 : Practical work

Sediment samples which have shown elevated Cd concentrations have tended to be found on the northern side of the Bank (Figs 1.2 and 1.3; Rowlatt and Lovell, 1994a). A seasonal (May – October) jet-like circulation described by Brown et al. (1999) forms a direct pathway for water adjacent to the north-east coast of England to pass offshore and around the northern flank of the Bank. Accordingly, it was decided to undertake a north-south transect crossing this pathway (Fig. 3.1) during times of high and low primary production, to see whether differences in sediment phytoplankton concentration might correlate with variable Cd concentrations.

A Scanfish section (towed undulating CTD; e.g. Fernand 1999) (fig. 5.1) performed from RV Corystes in August 2000 (CORY 11/00) illustrates the structure of the water column, locations of the jet-like circulation and the distribution of primary production (indicated by high fluorescence intensity, for example fig 5.1). Initially, 11 NIOZ cores were taken along this section at equally spaced intervals in order to select areas which would yield a range of chlorophyll concentrations. Following analysis of surface sediment for chlorophyll concentration three locations were selected for further sampling. Five samples were taken from a site in deep water, twenty from a site corresponding to peak sediment chlorophyll and corresponding near-bed water column production and five from a site in shallow water on top of the Bank. Core samples were preferred to grab samples, both to reduce the amount of low density surface material that might be lost on sampling and to enable undisturbed samples to be taken at depth.

The 30 stations at the three sites were repeated during January 2001 from RV Cirolana (CIRO 1/01), during a period of relatively low primary production (production on the Dogger Bank occurs all year round). The sediments were digested with HF and total metal content determined in spring 2001. Detailed methods of sediment preparation and analysis are presented in Annex 3. In addition to the work undertaken in this contract, data collected under a number of MAFF funded projects (AE1214 & AE1219) during 1999 were also analysed:

Water samples at three depths (surface, thermocline (“middle”) and bottom) were collected from a transect of CTD stations offshore from the north east coast of

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Fig 5.1a

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Fig 5.1b

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England (fig. 3.1) and analysed for cadmium. The aim of this was to look for relationships between solid (particulate) and aqueous phases in the water column.

A “dogleg” transect of NIOZ cores which cut across the Dogger Bank were collected (fig 3.1). In the laboratory, these were sliced and the <2 mm fraction of sediment from the surface 1 cm, 5 cm and 10 cm depths were digested with HF and analysed for total metals. Unfortunately, data from this work did not become available until late 2001.

5.3 : Analytical Quality Control

The sediment metal data in this report have come from different studies and have been analysed either by CEFAS or BGS. Core samples from 1998 and 1999 were digested at BGS and analysed at CEFAS; so quality control data are also presented for these samples. To investigate the comparability of the different datasets, data from the certified reference materials analysed by each laboratory have been collated, and the comparison shown in fig 5.2 (data are presented in Annex 4). Data are shown for more metals than are reported in the current study because they provide a useful reference. The bars show 95% confidence intervals (CI). There is generally very good agreement between data from the two laboratories, so the data can be used interchangeably. However, the poor copper data from CEFAS resulted from contamination during analysis. Low recoveries for Cr are consistent with previous experience, and reflect the difficulty in extracting this metal from the CRM sediment matrix by HF. The data generally show good reproducibility, with 95% CIs generally below 5% (Cd 95% CI 5-8%).

Seawater samples were taken in 1999, filtered at 0.45µm and preserved with 0.1% nitric acid. They were analysed for cadmium by WRc, using the seawater CRM CASS-3, which has a certified Cd concentration of 30 ng l-1 +/- 5 ng l-1. The mean concentration found was 32.8 ng l-1 (95% CI = 1.8) so the observed values fell within the expected range.

5.4 : Results

Fig 3.1 shows colour coded sample locations.

5.4.1 : Seasonal sediment metal concentrations

Full details of chlorophyll a and metal concentration data are presented in Annex 5 and summary statistics in Table 5.1.

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Fig 5.2

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Table 5.1 : Summary statistics from seasonal metal concentrations in sediments from Dogger 2000-01 transect.

Al% Cd Cr Cu Fe% Li Mn Ni Pb Rb V Znmg/kg mg/kg mg/kg mg/kg mg/kg Mg/kg mg/kg mg/kg Mg/kg mg/kg

Summermean 1.64 0.05 16.8 2.4 0.73 6.1 239 4.3 9.4 30.0 19.7 15.0n 20 2 20 20 20 20 20 20 20 20 20 20sd 0.46 * 7.72 0.97 0.22 2.97 101 2.50 3.91 9.64 5.79 6.88max 2.73 0.06 33.0 5.4 1.34 15.3 566 12.2 21.0 55.1 36.2 31.0min 1.26 0.05 7.7 1.4 0.46 4.2 149 2.6 6.8 22.7 12.4 7.4

Wintermean 1.61 0.05 17.4 2.4 0.74 5.8 248 4.1 9.1 28.9 19.4 12.1n 26 1 26 26 26 26 26 26 26 26 26 26sd 0.42 * 8.72 0.96 0.23 2.36 110 2.03 3.32 8.39 5.19 4.96max 2.53 * 39.0 4.1 1.39 11.4 561 8.8 16.4 48.4 32.6 23.2min 1.25 * 7.3 1.3 0.41 4.1 107 2.0 6.1 23.2 12.7 5.3

As shown in Table 5.1, similar ranges in metal concentrations were generally found during summer and winter. It is somewhat unfortunate that concentrations of the metal of primary interest here – cadmium – were all at or below the detection limit of 50 µg kg-1. This makes further investigation of mechanisms involved in cadmium cycling difficult. However, data from other determinands show some interesting patterns. Figs 5.3a-f (seasonal plots) illustrate data for Al, Cr, Fe, Mn, Pb and Zn across the transect. Metals which tend to follow the clay content, such as Al and Pb, show similar concentrations in samples collected in winter and summer. Metals which might be expected to be subject to variation according to redox conditions (oxygen availability), such as iron and manganese, show greater variability between the seasons. Perhaps surprisingly, chromium falls into the latter category – it would normally be expected to follow the behaviour of the clay minerals. Zinc concentrations are fairly consistent between the seasons.

The association of some metals with redox-dominated coatings on sediment at the Dogger Bank was found in earlier work (Whalley et al., 1997), although at that time more complex associations were found. If, as is generally accepted, most trace metal sorption is controlled by the clay content (represented by Al) or Fe/Mn oxy-hydroxide coatings (represented by Fe), then there ought to be linear relationships between these metals and trace metals. The relationships are shown in Fig 5.4a-c and 5.5a-c.

From Figs 5.3-5.5, it appears that there may be two populations of sediment type – the smaller group of 4 samples come from the most northerly part of the transect, in deep water (> 70 m) off the Bank. As those metals most strongly associated with the mineral fraction, i.e. Al, Li and Pb, all show similar behaviour in this subset, it may be concluded that these sediments are of a different matrix composition to those on the Bank. Chromium is quite scattered in the seasonal plot, but shows a relationship with iron or manganese (e.g. fig 5.5a Cr:Fe r2 = 0.774, n = 45). Where metals follow the redox-sensitive metals iron and manganese, it is likely that they are present primarily through adsorption on to Fe and Mn oxy - hydroxide coatings. Lead is strongly

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Fig 5.3a+b

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Fig 5.3c+d

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Fig 5.3e+f

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Fig 5.4

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Fig 5.5

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associated with Al (r2 = 0.967 n = 46). It shows similar behaviour to that of Fe and Mn, but two sample populations are evident if these metals are used as normalisers. Such relationships suggest that Pb behaviour is most strongly governed by the clay content, with the oxy-hydroxide coatings upon fines being of secondary importance. Zinc behaviour most closely resembles that of Al and is quite scattered when compared with Mn. Additionally for zinc, two sample populations seem to be present regardless of whether Al, Fe or Mn are used as normalisers.

Fig 5.6 shows chlorophyll a concentrations in the top 3 cm of the sediment (slices were taken at 0 - 0.5 cm, 0.5 – 1 cm, 1 – 2 cm and 2 – 3 cm depths). Fig 5.7 shows data taken from the same depths, where total pigment = chlorophyll a + pheopigment (decaying plant material). As expected, there were generally lower concentrations of chlorophyll a in the surface 1 cm of sediment during winter than in summer (fig 5.6a+b). No patterns are observed in the data shown in fig 5.6, although in the shallow waters on top of the Bank at the southernmost end of the sampling area, there was little difference in chlorophyll a concentrations between the seasons, suggesting that production was continuing in these fairly shallow waters even in winter.

Fig 5.7 show the measured total chlorophyll in the top 3cm of sediment from the Dogger 2000-01 samples. Patterns are not evident in the total pigment during summer, but in winter the top 1cm generally contains less plant matter than the 1-3 cm depths.

5.4.2 : Discussion of results from seasonal sediment sampling

The aim of this element of the work was to analyse metals and chlorophyll a in surface sediments from the Dogger Bank region, from sites which showed as wide a range in concentrations of chlorophyll a as possible. Samples were collected in July and January to try and identify any seasonal effects caused by variable inputs of phytoplankton-Cd. We expected to find higher sediment concentrations of Cd in samples collected in winter compared to those collected during summer, because primary production is lower in winter and there was likely to be less organic material in the sediment. The range in summer concentrations of chlorophyll a was 1.12 - 3.12 mg m-2, giving an expectation of variable Cd concentrations within the sediment. However, we could not further investigate the relationship between Cd and chlorophyll a in sediments as Cd concentrations were below the analytical detection limit of 50 ng g-1.

In a recently published paper, Stoeck and Kronke (2001) found low amounts of chlorophyll a in surface sediments at the Dogger Bank compared to further down the profile (4 and 8 cm). This contrasted with the German Bight, for example, where surface sediments were enriched in chlorophyll a when compared to those below the surface. The authors attributed these observations to surficial organisms at the Dogger ingesting chlorophyll a faster than those in deeper layers, and to advective transport of the organic matter into subsurface layers, which can occur under the physical conditions which are prevalent at the Dogger. Our data are therefore not inconsistent with cadmium being transported from the water column, via phytoplankton, to the sediments at the Dogger, but no direct association has been demonstrated.

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Fig 5.6

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Fig 5.7

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Other data from the transect showed that certain trace metals were associated with Fe or Mn-oxy-hydroxide coatings rather than with proxies for the clay content. Notably, chromium showed a dependence upon these coatings.

5.4.3 : Dissolved cadmium in waters off the north east coast and from the Dogger dogleg transect

Filtered water samples (<0.45 µm) had been collected opportunistically during CTD transects in 1999 as part of the MAFF contract AE1214. These samples were analysed to investigate whether any associations between dissolved and solid phase cadmium could be determined.

Cadmium analyses in seawater were performed using an MIBK extraction by Dr Sean Comber at WRc (Apte and Gunn, 1987). Figure 5.8a shows cadmium concentrations in samples from the Dogger dogleg transect and fig 5.8b those from the Tyne Tees transect. “B” represents samples taken from ~6 m from the seabed, “M” those from the thermocline (20-30 m depth), and “S” those from ~6 m below the surface of the water. Concentrations in samples were in the range 4-15 ng l-1. These are similar to the lower values found by Dr Rebekah Owens (pers. comm.) in samples collected in 1991 from RV Cirolana (two stations on Dogger Bank, Cd = 11 and 16 ng l-1; two at the mouth of the River Tyne, Cd = 11 and 69 ng l-1). There was little trend apparent in the dissolved phase across the transects, although there is some suggestion of slightly lower dissolved Cd in surface waters to the north of the Bank (Fig 5.8a). If this were genuine depletion in the surface waters, it would be mimicking the behaviour of dissolved phase nutrients.

There was no apparent relationship between dissolved cadmium (<0.45 µm) and chlorophyll a (fig 5.9a+b). Also, there was no observed relationship between dissolved cadmium and that in suspended particulates (fig 5.10a+b). Relationships with e.g. Fe or Mn in the particulate phase were also not apparent. Generally, a relationship between the dissolved and particulate phase would be expected because there is an equilibrium (the partition coefficient Kp) between these two phases. However, there may be difficulties with examining partitioning here, owing to the inherent variability in determining suspended load concentration within the water column. This meant that calculations were based on the absolute amount of particulate cadmium (µg) found in a known volume of water, rather than per mass of sediment (µg kg-1). This feature may have introduced artefacts into the Kp calculation (µg l-1 / µg kg-1). Other possible reasons for a lack of relationship are i) equilibrium has not been reached; ii) there is another influence (as yet unaccounted for) upon the equilibrium; iii) low concentrations led to noise disguising any relationship.

5.4.4 : Metals in cored sediments from the Dogger Bank dogleg, 1999

Sediment cores were collected using a NIOZ corer from the Dogger Bank region during summer 1999 (fig 3.1). The cores were frozen while on the ship and later transferred to the laboratory where they were defrosted and sliced at 0-1cm (“1cm”), 4.5-5.5 cm (“5 cm”) and 9.5-10.5 cm (“10 cm”). Samples were sieved at 2 mm, freeze-dried then HF-digested and analysed for metals. Figs 5.11a-h show metal profiles along the transect. Stations lying on the Bank are #160-167, with #156 lying to the north west end and #174 to the south east end of the transect.

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Fig 5.8

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Fig 5.9

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Fig 5.10

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The data for Al shown in fig 5.11a exhibit the change in sediment composition between the finer sediments off the Bank (higher Al content) and sandier sediments on the Bank. All the metals, excepting lead, show changes in concentration in samples collected from the Bank. Manganese concentrations in the samples show a marked increase over the Bank while those of aluminium and lithium decrease.

There is a linear relationship between Pb and Al (table 5.2 and fig 5.12b), suggesting that the lead is associated with the clay minerals as would be expected. However, a plot of Cr vs Al suggests that there are 2 populations of sediment types (Fig 5.12a) – samples with higher chromium concentrations are from the Bank. Usually, chromium is expected to be associated with the clay minerals, but here this metal shows alternative relationships (table 5.2 and fig 5.11g). A pattern with deeper samples from the Bank showing elevated concentrations is repeated for Cd, Cr, Fe, Mn and Zn (Figs 5.11 b, d, e, f and g).

Table 5.2 shows that there are relationships with trace metals between either metals in the clay fraction (e.g. Al) or with the redox-sensitive metals Fe and Mn. Zinc exhibits intermediate behaviour.

Table 5.2 : Metal / normaliser relationships in cored sediments

Normaliser (N)Aluminium Iron Manganese

Metal (y) r2 df significance r2 Df significance r2 df significance

Al ns 1 --- --- ns 1 --- ---Cd ns --- --- 0.637 47 ***** 0.613 46 *****Cr ns 1 --- --- 0.650 54 ***** 0.607 53 *****Fe ns 1 --- --- 0.641 53 *****Li 0.890 54 ***** ns 1 --- --- ns1 --- ---Mn ns 1 --- --- 0.641 53 *****Pb 0.836 53 ***** ns 1 --- --- ns1 --- ---Zn ns 1 --- --- 0.217 54 *** ns1 --- ---

Notesdf = degrees of freedomSignificance : P<0.001 = ***, P<0.0000 = *****ns = not significantns 1 = not significant with 2 populations of y:N evident from regression plot (for example, see fig 5.12a)

Investigating the relationships further shows that there are associations between eg Cd and Fe, Cd and Mn but not between Cd and Al (table 5.2 and fig 5.12d). These suggest an association between the redox-sensitive, oxy-hydroxide forming metals Fe and Mn and the trace metals. It is not yet clear why the surficial sediment does not tend to show elevated concentrations of these metals. In the case of cadmium, the strength of association with Fe or Mn in the sediments would suggest that whatever route brought it to the sediments, redox effects become dominant once the metal has been buried.

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Fig 5.11a+b

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Fig 5.11c+d

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Fig 5.11e+f

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Fig 5.11g+h

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Fig 5.12

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These data show different concentrations between sediments on the Dogger Bank compared to those north and south of the Bank. There is an apparent change in behaviour since clay-associated metals (Al, Li, Pb) generally show lower concentrations in sediments from the Bank while trace metals (Cd, Cr, Fe, Mn and Zn) generally show higher concentrations in sediments from the Bank. In particular, the difference in behaviour between clay-associated lead and chromium suggests an alternative source or pathway for chromium.

Fe or Mn (oxy-)hydroxides have a high affinity for sorption of dissolved metals (Whalley, 1995). These (oxy-)hydroxides can dissolve in oxygen-poor environments, such as might be expected in buried sediments, with concomitant release of sorbed metals. This can produce a cycling effect in surface sediments, where co-dissolved metals re-circulate up the sediment column until redox conditions become more oxidising, whereupon they re-precipitate. In the data reported here, a re-circulation effect may be occurring in the lower samples (5 cm and 10 cm depth) for the trace metals, assuming that dissolution of oxy-hydroxide coatings occurs further down the sediment column. It is unclear why the top 1 cm of sediment does not show these elevated concentrations.

Elevated concentrations in lower layers may be caused by mobility or dilution effects in the surface layer. It is unlikely that the greater concentrations at 5-10 cm depth represent the influence of any historical inputs, since sediments on the Dogger Bank are likely to be disturbed to 10 cm by storms on an annual basis. Therefore any historical differences in material being deposited would be expected to be distributed throughout the top 10 cm of sediment. An alternative possibility for metals being transported into the sediment is that raised by Stoeck and Kröncke (2001), where higher chlorophyll a concentrations were found in deeper sediments at the Dogger Bank. This was attributed partly to domination of the Dogger surface sediments by organisms ingesting chlorophyll a faster than those in deeper sediments, and partly to burial by advective transport of organic matter.

5.4.5 : Considering possible pathways for cadmium to reach sediments at the Bank.

Atmospheric metal inputs to the North Sea were reported in the OSPAR Quality Status Report for the Greater North Sea (OSPAR, 2000). Total atmospheric emission and deposition of cadmium was halved between 1987 and 1995, and atmospheric deposition accounted for about one third of the total cadmium input into the region, with >50% of the atmospheric inputs being received by the Southern Bight. The UK accounts for a large proportion of atmospheric emissions of cadmium by the Contracting Parties (59t out of 79t in 1995), so there may be a concern here. Using lead as a tracer for atmospheric inputs, the QSR 2000 reported that, in contrast to most other areas, lead concentrations in blue mussels from the Dogger Bank showed a significant downward trend. It is therefore difficult to identify current atmospheric sources of metals from the UK as leading to metal accumulation in Dogger sediments.

If phytoplankton do provide a pathway for low, dissolved concentrations of cadmium to be concentrated into the solid phase before deposition to the sediment, there is the possibility that nutrient/phytoplankton dynamics occurring on the flanks of the Dogger Bank may have a role in determining cadmium concentrations in sediments on the Bank. The local circulation pattern on the northern flank of the Bank is

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currently being investigated under the DEFRA-funded contracts AE12225 and AE1219 (“thin layers”). To date, experiments have not shown evidence for transport of dissolved or particulate phase material on to the Bank. Instead, decaying blooms appear either to be deposited approximately in situ, or to drift eastwards. It seems more likely that any process involving phytoplankton providing a pathway for cadmium into sediments at the Bank develops locally.

The observation that it is only the samples on the Bank which appear to have elevated concentrations of some metals such as Cd, is interesting given the existence of a seasonal transport pathway which runs past the northern flank of the Bank (Brown et al 1999). Currently, our data provide no evidence for the direct transport of metals from the coastal zone within the seasonal jet-like circulation to the Dogger Bank. The suggestion by Wang et al. (2001a+b) that eutrophication might provide enhancement of cadmium transfer from dissolved to solid phase has not been investigated here, but may be of interest with respect to indirect pathways.

Observations of the relationship between zinc-limited phytoplankton and cadmium uptake were studied in several laboratories in the early 1990s. More recently, Wang’s group have been investigating associations between phytoplankton and radiolabelled cadmium added to natural seawater in laboratory studies. We have not found other work on particulate phase associations between chlorophyll a and cadmium in the literature. The work reported here, which has been carried out under a number of DETR, MAFF and DEFRA-funded contracts, represents a large resource in both financial terms and in the range of staff skills required. This resource-intensity may reflect the lack of similar observations in the literature.

5.6 : Conclusions

Data from previous MAFF funded work (AE1214) had shown an association between cadmium and chlorophyll a concentration (a proxy for phytoplankton) in the waters off the north east English coast, extending out to the Dogger Bank. A literature review undertaken for this project revealed evidence that under laboratory conditions, phytoplankton can actively take up cadmium under certain conditions, such as when zinc concentrations are low.

Work was undertaken to explore the association between cadmium and chlorophyll in surface sediments in the Dogger region. Such a correlation would provide an explanation for previously recorded elevated cadmium concentrations in some sandy sediments from the Dogger Bank. Sediment core samples were taken from similar locations during periods of ‘high’ (summer) and ‘low’ (winter) productivity to identify whether variable phytoplankton inputs produced variable cadmium concentrations in the sediment.

Seasonal variation in chlorophyll a and cadmium concentrations were examined in 30 surface sediment samples from three locations in the Dogger Bank region. However, on this occasion, all the cadmium concentrations were below the detection limit of 0.05mg kg-1. Other trace metals showed relationships either with metals representing the clay fraction (Al and Li), or with those representing (oxy-)hydroxide coatings (Fe

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and Mn). Chromium was associated with Fe and Mn while lead was associated with Al and Li.

No associations were found between dissolved cadmium in the water column and chlorophyll a or metals in the particulate phase, such as cadmium, iron, in archived water samples from 1999.

Data for metals in sediment cores from a transect across the Bank performed in 1999 were analysed. These showed elevated concentrations cadmium (up to 0.34 mg kg -1 ) and chromium (up to 78 mg kg -1) in sediments from 5 cm and 10 cm depths on top of the Bank, but not in the surface (top 1 cm) sediment layer. Elevated concentrations were not apparent in the finer sediments on either side of the Bank.

Under oxidising conditions, iron and manganese can form (oxy-)hydroxide coatings on particles. These coatings, which adsorb trace metals, can dissolve when oxygen becomes limiting and release associated metals into solution. Those metals which were elevated at the 5 cm and 10 cm depths were associated with Fe and Mn, suggesting that (oxy-)hydroxide coatings were the primary control over the cycling of Cd and Cr in the sediment. Other workers (Stoeck and Kröncke, 2001) recently found higher chlorophyll a concentrations in deeper sediments at the Dogger Bank. They attributed burial of chlorophyll a partly to biological and partly to physical processes. These data are therefore not inconsistent with natural processes leading to elevated cadmium concentrations in the sediments, but fail to provide a clear explanation for the mechanism.

The data for elevated metal concentrations in sediments were considered against knowledge of the recently described seasonal, jet-like circulation which flows from the north east English coast past the northern flank of the Dogger Bank, as reported by Brown et al. (1999). No evidence for direct transport of contaminants could be found, since the elevated concentrations were apparent in deeper sediments across the whole of the Bank.

5.7 : Further work

Analysis of retained 5cm and 10cm sediment core samples from Dogger transect 2000-01 for metals, to allow assessment of cadmium concentrations at greater depth (Stoeck and Kronke, 2001).

Analysis of lead isotopes in suspended particulate samples from water column transects, to positively identify any transport of contaminants from the north east coast out to the Dogger Bank.

Investigation of possible indirect pathway for eutrophication to enhance transfer of cadmium from dissolved to particulate phase (Wang et al. 2001a+b).

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5.8 : References

Apte S.C. and Gunn A.M. (1987) Rapid determination of copper, nickel, lead and cadmium in small samples of estuarine water. Analytica Chimica Acta, 193, 147-156.

Brown J, Hill AE, Fernand L and Horsburgh KJ. (1999) “Observations of a Seasonal Jet-like Circulation at the Central North Sea Cold Pool Margin” Est Coast Shelf Sci 48 (3) 343-355

Brown J, Fernand L, Horsburgh KJ, Hill AE. and Read JW. (2001). “Paralytic shellfish poisoning on the east coast of the UK in relation to seasonal density-driven circulation” J Plankton Res 23 105-116

Fernand, L., 1999. High resolution observations of the velocity field and thermohaline structure of the western Irish Sea gyre. PhD Thesis, University of Wales, Bangor, 93pp.

OSPAR Commission. (2000) Quality Status Report 2000. Region II – Greater North Sea. OSPAR Commission, London. 0 946956 48 0

OSPAR Commission. (2000) Quality Status Report 2000. Region V – Wider Atlantic. OSPAR Commission, London. 0 946956

Stoecke T and Kröncke I (2001) “Influence of particle mixing on vertical profiles of chlorophyll a and bacterial biomass in sediments” Est Coast Shelf Sci 52 783-795

Rowlatt S and Lovell D. (1994a) Survey of contaminants in coastal sediments. MAFF, Burnham-on-Crouch. DoE research contract PECD 7/7/358

Wang W-X and Dei RC. (2001a) “Metal uptake in a coastal diatom influenced by major nutrients (N, P and Si)” Wat Res 35 (1) 315-321

Wang W-X and Dei RC. (2001b) “Effects of major nutrient additions on metal uptake in phytoplankton” Env Poll 111 233-240

Whalley C, Rowlatt S, Jones L, Bennett M and Campbell S. (1997) Metals in sediments and benthos from the Dogger Bank, North Sea. CEFAS Burnham-on-Crouch, DoE contract CW0 301

Acknowledgements

Thanks to MLL, DETR and DEFRA for funding this work. Also, thanks to John Thain and Bryn Jones (CEFAS), Jenny Cook (BGS), Sean Comber (WRc), Stephen George (University of Stirling) and the crews of the research vessels for their help.

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Metal Biological Activity Final Report

Annex 1 : Preparation of biological samples

Biota were collected on RV Cirolana 4B/00 in June 2000, and sub-sampled within 1 hour of capture. Dab (15-20cm) were killed with a cranial blow or their spine severed and the liver collected. Benthos samples were frozen at –20 º C for 20-30 minutes to anaesthetize them immediately prior to sample excision (see below). The samples for MT measurement were stored in cryovials at –80 º C.Asterias rubens – pyloric caecum was collected (Temara et al, 1997)Eupagurus – hepatopancreas collectedEchinocardium cordatum – gonads removed, intestines washed of sediment and collected. Tissue samples remaining after MT and protein analyses were archived at –80°C.

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Metal Biological Activity Final Report

Annex 2 : Analytical procedures for metallothionein and protein analyses

(Analyses performed by Prof Stephen George, University of Stirling)

A2.1 : SOP Sample preparation for Metallothionein assays.

Fish Liver.*

Solutions:Stock Tris-HCl 0.5M pH 7.4, 6.05g Trizma base diss. 50mls, adjusted with 1NHCl and made up to 100mls. (autoclave) …working 20mM Tris diluted daily. Note : record all weighings in study book

Procedure:1. Take 300mg tissue samples into 4ml tubes, add 2 . 7 ml ice cold tris buffer, homogenise for 15secs with Polytron rotating blade homogeniser (small probe, cooled in ice before). Stand tube in ice. * for invertebrate digestive gland add 1mM PMSF to homogenisation buffer.

2. Spin 200xg 1min. (Jouan 4°C) to remove bubbles and debris.

3. Transfer supernatant to microfuge tubes with syringe (2 per sample). Sample 20ul from middle of each tube (ie 40ul total) to multiwell plate using Rainin wide ended tips, add 110ul 0.25N NaOH, then 100ul 2.5% SDS. Seal and store at room temperature overnight for homogenate protein assay ( 5ul 1/2 diln.). Centrifuge supematants at 13,000 rpm for 15mins at 4°C (precooled rotor).

4. Transfer and pool supernatants in 4ml centrifuge tubes with syringe, sample 40ul microplate for PMS proteins add 110ul 0.25N NaOH, then 100ul 2.5% SDS. Seal and store at room temperature overnight for homogenate protein assay ( 5ul 1/2 diln.).

5. Heat treat 1ml sample of pooled supernatant in microfuge tube and follow MT assay procedure, archive remainder of PMS at –86°C.

Daily procedure:

Decontaminate any positive radioactive areas. A. Prepare 2 x 50mls diluted stock Tris. 1 on ice, 1 room temp. Thaw 2 aliquots HgB to room temp. Cool homogeniser. Ic e buckets. Book refrigerated Eppendorf 10-12., fit and precool rotor.

Swab centrifuge and work area and count swabs, bag radioactive waste.

Additional protocols: DC Lowry protein: MT CD satn. assay.

Safety: Note precautions for protease inhibitors in COSHH assessment.

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Metal Biological Activity Final Report

A2.2 : SOP Lowry DC Protein assay.

Using Biorad kit (500-0116) components.

Solutions:Diluent: 0.3N NaOH, 0.5N NaOH Detergent: 1.7% and 5% SDSBlank soln.: 0.1N NaOH, 1% SDS 1mM Tris store frozen in plate stripsStandards. Stock 2.0 mg / ml BSA in 0.1N NaOH (store frozen in aliquots)Make up in microtitre plate strip.Final conc. 0.1 0.2 0.4 0.6 0.8 1.0 1.2 1.4 mg /ml ul BSA 10 20 40 60 80 100 120 140ul 0.5N NaOH 38 36 32 28 24 20 16 12ul water 112 104 88 72 56 40 24 8 ul 5% SDS 40 ---------------------------------------------- Aliquot 100ul to multiwell strips, store frozen and thaw for each determination

Working reagent A’ : 20ul reagent S / ml reagent A. (stable 1week, if precipitates, warm and vortex to dissolve SDS). ( make up 3ml / plate )

Procedure:Set up plates,column 1 blank, 5ul 0.1N NaOH 1% SDS 1mM Triscolumns 2-4. 5ul standardscolumns 5-12 5ul unknowns 1-16 in quadruplicate

10% flounder liver homogenate in 20mM Tris, sample 40ul to uplate, add 110ul 0.25N NaOH, stand 15min, add 100ul 2.5% SDS, seal , mix and stand o/n room temp. Dilute 1/2 for assay (5ul).

add 25ul reagent A’ per well add 200ul reagent B per well

Labsystems Multiscan Genesis Protocol …DCprot. 9 mix 10secs. Stand 20 min, mix and read 690nm. Calib. Curve, calc. Can delete one outlier to give CV <5% if not, repeat.

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Metal Biological Activity Final Report

A 2.3 : SOP Metallothionein Cd saturation assay

Solutions

Stock Tris-HCl 0.5M pH 7.4 (autoclave)…working 20mM Tris diluted daily.Stock Cd, 1mg Cd/ml, = 125mg CdCl2 2 1/2 H20 / 50ml 20mM tris pH 7.4 HgB, 20mg/ml (2%) in 20mM Tris pH 7.4 (store 20ml aliquots frozen –15°C.)Radio Cd; 2 ug Cd and 1uCi Cd109 /ml.

100ul of 1mg/ml cold stock in 50ml 20mM tris pH 7.4 add 50uCi 109Cd.(store in lead castle room temp.)Record all weighings in study book.

Assay:

1. Pipette 700ul PMS or cytosol into 1.5ml microfuge vials and heat denature for 3min at 95°C. Cool in ice water. Centrifuge 10min at full speed Epp. microfuge .

2. Pipette up to 200ul sample (less made up to 200ul with 20mM tris (stock 1/25diln.)) in triplicate. ( for forth flounder 50ul 10% homog. PMS)

Backgr. = 200ul tris (3 per set of 21 samples)Total = 400ul 1/25 diln. tris (ie. 20mM) (3 per day)

3. Add 200ul radio Cd (use Eppendorf repette) to each tube , mix and stand for 15min at room temp.

Put Totals to one side.

4. Add 100ul HgB (eppendorf repette), mix, Heat denature for 3 min 95 °C (not total), cool in ice bath, stand for 2min

5. Centrifuge 10min full speed. ...put projections of lids to centre of rotor (ie hinge to outside).

6. Repeat of steps 4 and 5. Add 100ul HgB (eppendorf repette), mix, Heat denature 3 min 95 °C (not total), cool in ice bath, stand for 2minCentrifuge 10min full speed. ...put projections of lids to centre of rotor.

7. Carefully remove 100ul SN to counting vial taking great care not to suck up any loose denatured HgB.

LSC, Pico vials (Packard) and add 5ml scintillant, cap, label caps and count program 8 on Beta counter (C14 settings)GSC, LP2 tubes (ThermoLife sciences) count to 10K counts on Cobra gamma counter, 50-2000KeV.

Interpretation: HgB blank <8% total (if not, purchase new HgB), Exptl. Counts 10-50% of total (if not, adjust dilution or homogenate w/v if). Replicate variance <8%, one discard allowed (if not, repeat). Fro flounder use 150ul HTSN from 10% homogenate.Safety: See radioactive hazard assessment and COSHH assessments. Remember to fill in radioactivity log book.

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Annex 3 : Collection, preparation and analysis of sediment samples

Sample collection

Sediment samples were collected by NIOZ corer. For the biological effects work, sediment was collected from the surface 2-3 cm, avoiding the sides of the metal corer, and stored in plastic tubs and frozen at –20 ˚ C until they were prepared for analysis.

For the phytoplankton work, sub-samples were taken by inserting an acid-washed, 5cm diameter Perspex tube into the core. In samples from the 1999 CORY 5/99 cruise, sub-samples were later taken at 0-1cm, 4.5-5.5cm and 9.5-10.5 cm (where the core reached this depth). In samples from the cruises CORY 9A/00 and CIRO 1/01, one core was used to provide samples for chlorophyll a and porewater analysis, at depths of 0-0.5cm, 0.5-1cm, 1-2cm and 2-3cm. The sediment “disc” was split into two, and chlorophyll a analyses were performed on board the ship. Porewater analyses were performed in the laboratory by drying to a constant mass, which then allowed the chlorophyll a per mass of sediment to be determined. A further core was taken for metals determinations and frozen upright at –20 ˚ C, prior to sub-samples being taken at similar depths back in the laboratory.

Sediment preparation and metal analysis

In the laboratory, the cored sediments were defrosted and extruded, cutting slices with plastic spatulas at 0-1cm, 1-2cm, 2-3cm and 3-4cm and storing these in plastic petri dishes. The 0-1cm slice was taken forward for analysis by wet sieving at 2mm, freeze-drying and milling to a fine powder. The powder was stored in a self-sealing plastic bag. A 1cm thick slice was taken for metals analysis to provide sufficient sediment for milling and representivity: since the chlorophyll a had been determined on 0-0.5cm and 0.5-1cm sediment slices, a mean value was derived with which to compare the metal and chlorophyll a concentrations. Sediments for the metallothionein work were defrosted and wet sieved at 2mm before freeze-drying and milling in agate.

The total digestion and analysis of sediments for this contract was performed by CEFAS, BGS or BGS/CEFAS where BGS digested the samples and CEFAS analysed them. The CEFAS total digestion procedure uses a mixture of HF/HNO3 and applied microwave heating to closed vessels, prior to the addition of boric acid to neutralise the HF toxicity. Analysis was by ICP-MS (As, Cd, Cr, Cu, Li, Ni, Pb, Rb, V, Zn) and Flame AAS (Al, Fe and Mn). BGS digested sediments on a hotplate using a mixture of HF/HClO4/HNO3. Analysis was performed by ICP-AES for Al, Cr, Fe, Li, Mn, Ni, V and Zn, and ICP-MS for As, Cd, Cu, Pb and Rb. The data for total metals in sediments produced by BGS and CEFAS are comparable (Table A4.1). Quality control data are reported in the main text: 1 blank and 1 marine sediment certified reference material were carried out to 10 samples: the methods used have been taken through QUASIMEME inter-comparisons.

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Annex 4 : Recoveries from certified reference materials (CRMs)

Table A4.1 : Percentage recoveries from marine sediment CRMs BCSS-1 and MESS-2 by BGS or CEFAS

Data for CRMs Al As Cd Cr Cu Fe Pb Mn Ni V Zn

BGS mean % recovery 97.18 97.66 100.91 80.85 94.82 94.90 101.57 95.46 96.79 93.06 85.07n 13 13 13 13 13 13 13 13 13 13 1395% CI 1.55 6.36 7.98 0.93 3.86 1.27 4.89 2.23 1.74 1.87 0.63

CEFAS mean % rec 102.11 109.63 102.24 77.97 121.40 100.38 96.40 98.05 97.47 98.41 99.20n 48 48 51 48 49 48 48 48 48 47 4895%CI 3.62 6.77 5.29 3.35 10.31 3.43 3.31 3.91 3.38 3.26 5.66

Table A4.2 : Percentage recoveries from CRMs in samples which were digested by BGS and analysed by CEFAS

BGS digested and Al As Cd Cr Cu Fe Pb Mn Ni V ZnCEFAS analysed

mean % recovery 98.03 85.81 106.31 82.69 104.73 97.87 96.09 90.95 94.21 90.47 83.14Sd 17.27 12.75 14.43 5.70 15.81 12.93 5.87 9.79 3.68 21.52 9.99N 16 16 16 16 16 16 16 16 16 16 1695% CI 8.46 6.25 7.07 2.79 7.74 6.34 2.87 4.80 1.80 10.54 4.89

Table A4.3 : Absolute recoveries from BCSS-1 for non-certified Li and Rb content

BCSS Li RbBGS CEFAS BGS+CEFAS BGS CEFAS BGS+CEFAS

mean 40.70 40.54 38.0 79.14 69.25 73.4sd 0.64 3.50 1.4 8.93 6.76 2.8n 7 40 7 7 40 795% CI 0.47 1.09 1.05 6.61 2.09 2.09

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Annex 5 : Sediment metal and chlorophyll concentrations, Dogger Bank 2000-01sample code Al Mn Fe V Zn Cu Li Ni Cr As Rb Cd Pb cruise stn de

gmin lat declat N/

Sdeg min

longDeclong

E/W

CHL PHEO nominal equivalent

% mg/kg % mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg lat long mg / m2 mg / m2 position stnCW/00/29 1.26 175 0.46 12 9 1 4 3 8 2 24 <0.05 7 CORY 9A/00 110 55 2.74 55.046 0 2 41.98 2.700 0 2.46 9.24 4 44/45CW/00/22 1.44 212 0.61 16 14 2 5 3 15 3 27 <0.05 8 CORY 9A/00 102 55 17.414 55.290 0 2 33.214 2.554 0 1.43 7.67 5 46CW/00/21 1.40 186 0.58 17 9 2 5 3 8 3 26 <0.05 7 CORY 9A/00 101 55 17.416 55.290 0 2 32.766 2.546 0 1.82 13.27 6 47CW/00/13 1.38 206 0.61 16 12 2 5 3 12 3 25 <0.05 8 CORY 9A/00 92 55 17.42 55.290 0 2 32.26 2.538 0 2.24 14.02 7 48CW/00/11 1.44 380 0.95 21 13 2 5 4 31 3 24 <0.05 8 CORY 9A/00 90 55 17.149 55.286 0 2 31.794 2.530 0 2.43 10.85 8 52CW/00/24 1.38 178 0.55 15 9 2 4 3 8 4 26 <0.05 7 CORY 9A/00 104 55 16.847 55.281 0 2 33.231 2.554 0 1.81 10.48 9 54CW/00/19 1.47 275 0.79 19 11 2 5 3 18 4 23 <0.05 7 CORY 9A/00 99 55 16.804 55.280 0 2 32.81 2.547 0 2.67 17.74 10 55CW/00/14 1.41 232 0.67 18 11 2 5 3 14 4 26 <0.05 7 CORY 9A/00 94 55 16.86 55.281 0 2 32.22 2.537 0 2.34 17.73 11 56CW/00/10 1.38 174 0.52 15 9 2 5 3 11 4 27 <0.05 7 CORY 9A/00 89 55 16.82 55.280 0 2 31.76 2.529 0 1.89 9.90 12 57CW/00/25 1.39 156 0.50 14 10 2 5 3 8 3 26 <0.05 7 CORY 9A/00 105 55 16.606 55.277 0 2 33.209 2.553 0 2.17 13.97 13 58CW/00/17 1.33 149 0.47 14 7 2 4 3 8 4 26 <0.05 7 CORY 9A/00 98 55 16.595 55.277 0 2 32.782 2.546 0 1.64 10.09 14 59CW/00/15 1.46 326 0.82 20 21 2 5 3 22 3 25 <0.05 8 CORY 9A/00 95 55 16.584 55.276 0 2 32.255 2.538 0 2.44 20.96 15 60CW/00/26 1.69 566 1.34 29 17 3 5 5 33 3 26 <0.05 10 CORY 9A/00 106 55 16.32 55.272 0 2 33.195 2.553 0 1.33 9.63 16 62CW/00/18 1.41 228 0.69 18 14 2 5 3 12 4 26 <0.05 8 CORY 9A/00 97 55 16.305 55.272 0 2 32.686 2.545 0 2.34 12.47 17 63CW/00/16 1.55 270 0.78 21 31 3 6 4 19 3 27 <0.05 9 CORY 9A/00 96 55 16.318 55.272 0 2 32.164 2.536 0 1.12 10.54 18 64CW/00/08 1.43 337 0.86 20 11 2 5 3 20 3 24 <0.05 8 CORY 9A/00 87 55 16.25 55.271 0 2 31.82 2.530 0 2.06 11.66 19 65

55 17.417 55.290 0 2 31.735 2.529 0 2.23 10.96 20 66CW/00/31 2.39 182 0.80 26 20 4 11 8 19 5 46 <0.05 16 CORY 9A/00 115 55 46.707 55.778 0 2 12.295 2.205 0 3.12 19.66 21 67CW/00/32 2.48 174 0.79 25 20 3 10 8 19 5 47 <0.05 15 CORY 9A/00 116 55 46.687 55.778 0 2 12.999 2.217 0 2.47 19.85 22 68CW/00/30 2.73 195 1.08 36 28 5 15 12 28 7 55 0.06 21 CORY 9A/00 114 55 47.069 55.784 0 2 12.307 2.205 0 2.55 17.74 23 69/70CW/00/33 2.44 172 0.74 23 24 3 10 7 23 4 45 0.05 14 CORY 9A/00 117 55 46.887 55.781 0 2 12.605 2.210 0 2.89 20.35 24 71CW/01/06 1.25 185 0.52 13 7 1 4 2 8 <2 23 <0.05 6 CIRO 01/01 41 55 2.651 55.044 0 2 41.376 2.690 0 2.57 4.97 1 107CW/01/07 1.31 236 0.65 16 10 2 4 3 25 3 24 <0.05 7 CIRO 01/01 42 55 2.556 55.043 0 2 41.79 2.697 0 2.70 7.69 2 109CW/01/08 1.33 197 0.55 15 8 1 4 3 10 2 24 <0.05 6 CIRO 01/01 43 55 2.476 55.041 0 2 42.077 2.701 0 3.5 6.5 3 111CW/01/09 1.31 232 0.64 16 8 2 5 3 16 3 25 <0.05 7 CIRO 01/01 44 55 2.722 55.045 0 2 41.697 2.695 0 3.23 4.66 4 110CW/01/11 1.50 399 1.09 25 15 4 5 4 28 <2 25 <0.05 9 CIRO 01/01 46 55 17.356 55.289 0 2 33.329 2.555 0 0.64 6.07 5 102CW/01/12 1.39 178 0.58 16 11 2 5 3 11 3 25 <0.05 7 CIRO 01/01 47 55 17.433 55.291 0 2 32.697 2.545 0 0.80 8.72 6 101CW/01/13 1.37 215 0.56 15 9 2 5 3 12 3 25 <0.05 7 CIRO 01/01 48 55 17.421 55.290 0 2 32.226 2.537 0 0.64 7.25 7 92CW/01/16 1.38 187 0.57 15 9 2 5 3 10 3 25 <0.05 7 CIRO 01/01 52 55 17.09 55.285 0 2 31.788 2.530 0 0.51 6.63 8 90CW/01/18 1.52 342 0.91 21 11 2 5 3 14 3 25 <0.05 8 CIRO 01/01 54 55 16.737 55.279 0 2 33.243 2.554 0 0.74 6.00 9 104CW/01/19 1.37 132 0.47 13 8 2 4 2 7 3 26 <0.05 7 CIRO 01/01 55 55 16.888 55.281 0 2 32.823 2.547 0 0.61 7.63 10 99CW/01/20 1.29 107 0.41 13 5 1 5 2 8 4 25 <0.05 7 CIRO 01/01 56 55 16.865 55.281 0 2 32.253 2.538 0 0.48 6.68 11 94CW/01/21 1.48 339 0.90 21 12 2 5 3 22 3 25 <0.05 8 CIRO 01/01 57 55 16.83 55.281 0 2 31.76 2.529 0 0.99 6.51 12 89CW/01/22 1.54 292 0.86 20 11 2 5 4 24 4 26 <0.05 8 CIRO 01/01 58 55 16.668 55.278 0 2 33.197 2.553 0 0.51 8.21 13 105CW/01/23 1.42 258 0.72 19 10 2 5 3 11 4 25 <0.05 8 CIRO 01/01 59 55 16.569 55.276 0 2 32.74 2.546 0 0.58 6.59 14 98CW/01/24 1.39 331 0.83 20 9 2 5 4 18 3 23 <0.05 8 CIRO 01/01 60 55 16.728 55.279 0 2 32.227 2.537 0 0.51 5.88 15 95CW/01/26 1.50 280 0.75 19 11 2 5 5 13 3 25 <0.05 8 CIRO 01/01 62 55 16.363 55.273 0 2 33.14 2.552 0 0.84 7.30 16 106CW/01/27 1.68 561 1.39 33 19 4 6 5 39 3 26 0.05 11 CIRO 01/01 63 55 16.287 55.271 0 2 32.675 2.545 0 0.57 5.97 17 97CW/01/28 1.55 422 1.05 24 15 3 5 3 18 4 25 <0.05 9 CIRO 01/01 64 55 16.334 55.272 0 2 32.124 2.535 0 0.58 5.83 18 96CW/01/29 1.33 134 0.46 13 8 1 4 2 8 3 26 <0.05 7 CIRO 01/01 65 55 16.214 55.270 0 2 31.893 2.532 0 0.71 7.09 19 87CW/01/30 1.48 397 1.01 21 14 2 5 4 39 3 24 <0.05 8 CIRO 01/01 66 55 17.368 55.289 0 2 31.735 2.529 0 0.77 5.85 20 91CW/01/31 2.36 148 0.76 23 18 4 10 8 18 4 44 <0.05 15 CIRO 01/01 67 55 47.321 55.789 0 2 12.358 2.206 0 1.08 13.68 21 115CW/01/32 2.47 204 0.81 25 19 4 10 8 23 5 46 <0.05 15 CIRO 01/01 68 55 47.218 55.787 0 2 12.374 2.206 0 0.86 16.37 22 116CW/01/33 2.45 142 0.78 25 19 4 11 8 20 3 46 <0.05 16 CIRO 01/01 69 55 47.132 55.786 0 2 12.428 2.207 0 1.09 18.03 23 114CW/01/35 2.53 177 0.85 27 23 4 11 9 19 4 48 <0.05 16 CIRO 01/01 71 55 46.944 55.782 0 2 12.566 2.209 0 0.64 17.31 24 117CW/01/10 1.29 187 0.52 14 6 2 5 2 9 3 25 <0.05 6 CIRO 01/01 45 55 2.654 55.044 0 2 41.902 2.698 0 2.43 12.28 4 107CW/01/34 2.36 180 0.73 23 19 4 9 6 23 4 44 <0.05 14 CIRO 01/01 70 55 47.041 55.784 0 2 12.482 2.208 0 1.17 17.67 23 114