S C Johannessen
email: johannessen@pac.dfo-mpo.gc.ca

R W Macdonald

C A Wright

B Burd

D P Shaw

A Van Roodselaar

gia sediments

Polychlorinated biphenyls (PCBs) are relict contaminants, while polybrominated diphenyl ethers (PBDEs) are in increasing use. Using sediment cores collected in the Strait of Georgia, we demonstrate that the surface sediment concentration of PCBs is largely determined by environmental processes, such as sediment accumulation and mixing rates, while that of PBDEs is strongly influenced by proximity to source. The Iona Island wastewater outfall appears to be a primary pathway for PBDEs. As well, Vancouver Harbour is highly contaminated with both classes of chemical. BDE-209, the main component of deca-BDE, is the dominant PBDE congener. Environmental debromination is not evident. Currently, the ranges of the surface concentration of PCBs and PBDEs are similar to one another, but that will change in the future, as the concentration of PBDEs continues to rise. The experience with PCBs suggests that if PBDEs were banned today, it would take decades for inorganic sediment to bury them. PCB, PBDE, contamination history, sediment, cores, Strait of Georgia

Introduction

Polychlorinated biphenyls (PCBs) and polybrominated diphenyl ethers (PBDEs) are persistent, toxic, bioaccumulative, manufactured chemicals that are widely distributed in the environment, but their histories of use are different [START_REF] Alaee | Recent progress in understanding of the levels, trends, fate and effects of BFRs in the environment[END_REF][START_REF] De Wit | An overview of brominated flame retardants in the environment[END_REF][START_REF] Rapaport | Historical atmospheric inputs of high molecular weight chlorinated hydrocarbons to eastern North America[END_REF]. Introduced in the 1930s as lubricants and liquid insulators and later used for a wide variety of industrial and household uses, PCBs (estimated total global usage of 1.2 Mt, [START_REF] Macdonald | Contaminants in the Canadian Arctic: 5 years of progress in understanding sources occurrence and pathways[END_REF] were banned in stages in the late 1970s and early 1980s as a result of concerns over their effects on the environment and on human health (Fig. 1). Despite their having been banned about 30 years ago, PCBs continue to cycle in marine and terrestrial ecosystems, where they still present health threats, especially to high trophic level, long-lived animals like killer whales [START_REF] Hickie | Killer whales (Orcinus orca) face protracted health risks associated with lifetime exposure to PCBs[END_REF][START_REF] Ross | High PCB concentrations in free-ranging Pacific killer whales, Orcinus orca: Effects of age, sex and dietary preference[END_REF].

PBDEs are more recent arrivals, used primarily as flame retardants on household goods, including furniture and electronics. Their production was estimated at 40 kt/yr in 1990 and had grown to about 70 kt/yr by the late 1990s [START_REF] Wania | Assessing the long-range transport potential of polybrominated diphenyl ethers: A comparison of four multimedia models[END_REF]. PBDEs appeared in the marine environment in the late 1970s and have been increasing in concentration ever since. They have been measured in sediment, water, and marine organisms (for example [START_REF] De Boer | Do flame retardants threaten ocean life?[END_REF][START_REF] Elliott | Polybrominated diphenyl ether trends in eggs of marine and freshwater birds from British Columbia, Canada, 1979-2002[END_REF][START_REF] Ikonomou | Spatio-temporal and species-specific variation in PBDE levels / patterns in British Columbia' s coastal waters[END_REF][START_REF] Ross | Fireproof killer whales (Orcinus orca): flame-retardant chemicals and the conservation imperative in the charismatic icon of British Columbia, Canada[END_REF], where they pose a threat as disrupters of the thyroid hormone system, which controls the development of the central nervous system [START_REF] Martin | An Asian quandary: where have all of the PBDEs gone?[END_REF]. PBDEs have been used in three market formulations: penta, octa and deca. The penta formulation, which contains the lightest congeners, was voluntarily phased out of use in Canada in the early 2000s, and has recently been banned in Canada, the United States and parts of Europe. The octa formulation has not been used locally. However, the heaviest industrial formulation, deca-BDE, remains on the market. PCBs and PBDEs are particlereactive, especially the heavier PBDE congeners [START_REF] De Boer | Polybrominated diphenyl ethers in influents, suspended particulate matter, sediments, sewage treatment plant and effluents and biota from the Netherlands[END_REF], so marine sediments serve as an archive of the fate of these chemicals and as a proxy for their historical discharge to the marine environment.

The Strait of Georgia (Fig. 2), a large inland sea on the west coast of British Columbia, Canada, has a rapidly growing population of ~4 million, almost half of whom live in Greater Vancouver, at the mouth of the Fraser River. The Strait has long been valued for fishing, transportation and recreation, but, like most coastal seas, it has also been used directly and indirectly for discharge of industrial waste, including pulp mill, mining and municipal wastewater effluents. Sediments in the Strait of Georgia record the entry of a wide variety of anthropogenic contaminants, including lead [START_REF] Macdonald | Accumulation of heavy metals (Pb, Zn, Cu, Cd), carbon and nitrogen in sediments from Strait of Georgia, B.C., Canada[END_REF], mercury [START_REF] Johannessen | Historical trends in mercury sedimentation and mixing in the Strait of Georgia, Canada[END_REF], chlorinated dioxins and furans (PCDD/Fs) and PCB-77 [START_REF] Macdonald | A history of octachlordibenzo-p-dioxin, 2,3,7,8-tetrachlorodibenzofuran, and 3,3' ,4,4'tetrachlorobiphenyl contamination in Howe Sound[END_REF], polynuclear aromatic hydrocarbons (PAHs, [START_REF] Yunker | Alkane and PAH depositional history, sources and fluxes in sediments from the Fraser River Basin and Strait of Georgia[END_REF] and industrial detergents [START_REF] Shang | Persistence of nonylphenol ethoxylate surfactants and their primary degradation products in sediments from near a municipal outfall in the Strait of Georgia, British Columbia, Canada[END_REF].

Here we examine dated sediment cores collected from the Strait of Georgia which have been analyzed for PCB and PBDE congeners. We use these sedimentary records in comparison with other records, including sediment traps and effluent analyses, to infer the controls on the distributions of PCBs and PBDEs in the sediments, to seek evidence for debromination of deca-BDE to lighter, more toxic congeners, and to assess the present role of the Iona Island wastewater outfall, which discharges directly into the Strait, in the delivery of these chemicals to the Strait of Georgia.

Methods

Sediment core collection and dating

Seven sediment cores were collected in 2003 and 2004 from the central and northern Strait of Georgia (Fig. 2), using a Pouliot box corer (20 cm × 30 cm cross section). The cores were all approximately 50 cm long and composed mainly of silty mud.

Immediately on recovery, the cores were sectioned for analysis into 1 cm intervals for the top 10 cm, 2 cm intervals for the next 10 cm and 5 cm intervals for the remainder of the core. A subsample from each depth interval was analyzed by Flett Research Ltd., Winnipeg, Canada, for 210 Pb and 226 Ra to be used for radio-dating. 210 Pb was measured in all sections of each core (~22 samples / core), following the procedure of [START_REF] Eakins | A new procedure for the determination of lead-210 in lake and marine sediments[END_REF], in which the activity of 210 Pb is inferred from that of 210 Po (counting errors < 3%). The activity of supported 210 Pb was determined as the average of the 226 Ra activity measured at three depths (top, middle, bottom) in each core, from the ingrowth of 222 Rn over at least 4 days.

Since the bottom waters of the Strait of Georgia are oxygenated, there is an active benthic community, and the surface sediments are bioturbated. The depth of the surface mixed layer in each core was determined by eye as the inflection point in the depth profile of the natural log of 210 Pb (Fig. 3). The sedimentation and mixing rates were determined assuming a constant supply of 210 Pb and a constant sedimentation rate [START_REF] Robbins | Geochemical and geophysical applications of radioactive lead[END_REF], with mixing described by advective-diffusive equations [START_REF] Lavelle | Accumulation rates of recent sediments in Puget Sound, Washington[END_REF].

Sediment traps

Moorings with two sediment traps each (attached 20 and 50 m off the bottom) were set out at stations GVRD-A1 and GVRD-B1 (Fig. 2) in October 2003 and retrieved in April 2004. Each trap collected 10 sequential samples that represented 21 days each. For each trap, we took a subsample from each of the 10 sequential vials, homogenized the subsamples and sent them for analysis of PCBs and PBDES, as described in section 2.4.

The data from the traps were used for comparison of the congener patterns with those measured in the sediment cores.

Carbon and nitrogen analysis

Carbon and nitrogen were analysed in the sediment core samples at the University of British Columbia, using the method of [START_REF] Calvert | On the organic carbon maximum on the continental slope of the eastern Arabian Sea[END_REF], in which the concentration of organic carbon is calculated as the difference between the concentrations of total and carbonate carbon. Total carbon is measured by combustion and gas chromatography in a CHS analyzer and carbonate carbon by coulometry (precision ± 1.6%, 1 standard deviation).

Analysis of PCBs and PBDEs

A full suite of 209 PCB congeners, with some co-elutions, and 40 PBDE congeners were measured by AXYS Analytical Services Ltd, in Sidney, B.C.. Both the PCBs and the PBDEs were measured by high resolution gas chromatography / high resolution mass spectroscopy (general precision ± 30%; precision of duplicates in a batch ± 20% of the mean). The PCB analysis was conducted according to the United States Environmental Protection Agency method EPA 1668A (EPA, 2003a), the PBDE analysis, according to method EPA 1614 (EPA, 2003b).

Calculation of 10-year inventories of PCBs and PBDEs in sediment cores

We calculated inventories of PCBs and PBDEs in the sediment cores for comparison with the surface concentration values. Unfortunately, the sedimentation rates at four of the stations (GVRD-3, -4, -5 and -6) were so high that the 50-cm core records were not long enough to capture the entire PCB record, which commenced in ~1930. Accordingly, we could not calculate a complete inventory for every core and have, instead, compared the inventories over a shorter period. Benthic mixing makes it inappropriate to assign a discrete date to any particular depth in these cores (See Guinasso and Schink, 1975 for explanation of the age-smearing effect of benthic mixing). We estimated a 10-year inventory for each core using a core depth that was scaled according to the sedimentation rate (sedimentation velocity at each site (cm/yr) multiplied by 10 years). In general, this method overestimates the inventory of the last 10 years of PCB deposition, since the concentration of that contaminant is decreasing, and the benthic organisms mix older, more highly contaminated material into the surface layer. Conversely, it underestimates the PBDE inventory because of the mixing of older, cleaner sediment into the surface layer. However, it provides a reasonable basis for comparison among the cores, since the biases are in the same direction for all cores.

Modelling historical deposition of PCBs and PBDEs

Bioturbation of surface sediment modifies the profile shape of a transient contaminant as it passes through this mixed layer (e.g. [START_REF] Guinasso | Quantitative estimates of biological mixing rates in abyssal sediments[END_REF]. We therefore modelled the historical deposition of PCBs in the cores, using a program developed in Matlab TM [START_REF] Johannessen | Historical trends in mercury sedimentation and mixing in the Strait of Georgia, Canada[END_REF] that explicitly accounts for the sedimentation and mixing rates. The method is most robust where there is an abrupt change, however, and there are too many degrees of freedom in the model to describe uniquely a flux that varies gradually, as those of both PCBs and PBDEs have done in most of these cores. However, at station GVRD-3, there was an abrupt change in the PCB concentration flux (as discussed in Section 3), and the model illustrates the effect of the newly-constructed outfall on the concentration of PCBs in the incident surface sediment. We have also used the model fit to some profiles (in Section 3) for a graphic illustration of the main trends with time.

Results and discussion

The concentration profiles in the sediments clearly reflect the known histories of PCB and PBDE discharge (Figs. 4 and5). PCBs appeared in ~ 1930, increased rapidly until the late 1960s, and then decreased. PBDEs first entered the local sediments in ~1978 and have been increasing rapidly ever since. Because of the relatively high sedimentation rates at stations GVRD-3, -4, -5 and -6, the deepest levels in those cores do not predate the entry of PCB into the environment ( 1930), but the recent PCB decline is reflected in every core. (The sedimentation rate at station GVRD-4, near the mouth of the Fraser River, was so high that we could not determine it with 210 Pb in a 50-cm core. Since we were unable to determine the amount of time represented by the core, we only measured contaminants at a few depths. Consequently, core GVRD-4 only appears on plots showing surface concentrations and not depth profiles or fluxes.) In contrast, PBDE profiles imply a pervasive rapid increase toward the sediment surface. Already in some parts of the world, the flux of PBDEs has overtaken that of PCBs. For example, the atmospheric flux of PBDEs to the Baltic Sea is estimated at almost 40 times that of PCBs [START_REF] Law | Levels and trends of brominated flame retardants in the European environment[END_REF].

The range of Total PCB concentration values in surface sediments (0 -1 cm depth, or 1 -2 cm in core GVRD-4) is similar to that of Total PBDEs (Table 2), but the two classes of chemical do not have the same spatial distribution at the surface (Fig. 6a). The has the highest surface concentration of PCBs, for example, the 10-year inventory of PCBs at that station is relatively low (Table 2, Fig. 6). The highest inventories of PCBs are found at stations GVRD-6 and GVRD-3 in the southern Strait, which have the highest sediment accumulation rates of the set. Station GVRD-4, near the mouth of the Fraser River, probably has the very highest inventory, but we were unable to calculate it, because, as mentioned above, the sediment accumulation rate at that site was too high.

The spatial pattern in the 10-year inventory of PBDEs is similar to that of the PCBs (Fig. 6), except that the inventory of PBDEs is very high at station GVRD-3, near the wastewater outfall. Excluding PBDEs at that station, the inventories of both classes of contaminants are strongly positively related to the sediment accumulation rate (Fig. 7).

However, the surface concentration is not. There is no significant correlation between surface concentration of PCBs or PBDEs and sediment accumulation rate (data not shown).

The measured surface concentration of a contaminant is strongly influenced by the local sediment accumulation and mixing rates (e.g. [START_REF] Johannessen | Historical trends in mercury sedimentation and mixing in the Strait of Georgia, Canada[END_REF]. The sediment accumulation rate is much lower in the northern Strait of Georgia than near the mouth of the Fraser River, which supplies about 80% of the sediment to the Strait [START_REF] Luternauer | Substrates of the Strait of Georgia, British Columbia[END_REF]. We infer that the concentration of PCBs in surface sediment is lower at the southern stations near the mouth of the Fraser River than in the northern Strait, because the inorganic particles discharged near these southern stations tends to dilute the sediment contaminant burden.

Benthic mixing compounds the effect of sedimentation rate on the observed surface concentration of PCBs. In the more recent (shallow) sections of the cores, the PCB concentration decreases toward the sediment surface because the discharge and flux to sediments was at its highest during the 1960s and 70s. When organisms mix the surface sediments, they tend to move the more contaminated, older sediment up into the less contaminated, post-1970 deposits. In an area like the northern Strait, which has a low sediment accumulation rate, the sediment that is mixed to the surface dates more closely to the PCB maximum and is, therefore, more contaminated than that mixed to the surface in an area with a higher sedimentation rate.

Because the PBDE concentration increases toward the surface, mixing has the opposite effect on its surface value than on that of PCBs. Older sediment mixed up to the surface has a lower concentration of PBDEs than the recent material, so the concentration of PBDEs measured at the surface of the sediments is lower than that in the particles currently accumulating. However, the surface concentration of PBDEs shows no relationship with sediment accumulation rate, even if it is normalized to the mixing rate (data not shown). Because of the recency of their arrival in this environment, PBDEs are still mainly concentrated near their sources. The linear relationship between sediment accumulation rate and the 10-year inventory of PBDEs shown in Fig. 7 implies an average concentration of PBDEs on particles deposited over the last ten years that is the same throughout the study area (~ 660 pg/g, from the slope of the regression line), except at the outfall station.

A study of the concentrations of PCBs and PBDEs in fish from remote European mountain lakes [START_REF] Gallego | Altitudinal gradients of PBDEs and PCBs in fish from European high mountain lakes[END_REF] similarly illustrated the importance of both history of emissions and environmental distribution processes. [START_REF] Gallego | Altitudinal gradients of PBDEs and PCBs in fish from European high mountain lakes[END_REF] showed that the concentration gradients of PCBS and PBDEs were directly related to temperature and to one another in the Pyrenees mountains, near long-standing sources of PBDEs, while, in the Tatras mountains, near newer sources of PBDEs, PCBs were still related to temperature, but PBDEs were not. They inferred from these observations that PCBs had had much longer to be redistributed by environmental processes like seasonal temperature cycling. Similarly, Spanish coastal sediments [START_REF] Eljarrat | Occurrence of polybrominated diphenylethers, polychlorinated dibenzo-p-dioxins, dibenzofurans and biphenyls in coastal sediments from Spain[END_REF] show correlations between PCBs and dioxin-like compounds but not PBDEs, except near the mouth of the river, which is likely a current pathway of PBDEs into that environment.

In the Strait of Georgia, the concentrations of PBDEs are highest near the wastewater outfall (GVRD-3, Table 2, Fig. 2). Work in other environments has similarly shown that the highest concentrations of PBDEs in sediment occur near industrial (particularly textile industries) and domestic wastewater discharges (e.g. [START_REF] Gevao | Spatial distribution of polybrominated diphenyl ethers in coastal marine sediments receiving industrial and municipal effluents in Kuwait[END_REF][START_REF] Law | Levels and trends of brominated flame retardants in the European environment[END_REF][START_REF] Samara | Determination of potential sources of PCBs and PBDEs in sediments of the Niagara River[END_REF] In contrast to PBDEs, PCBs are not actively used in homes or industry in Canada any more, although there is still relict PCB-containing equipment left, such as some transformers (Environment Canada, 2007, http://www.ec.gc.ca/wmddgd/default.asp?lang=En&n=13426772-1), and they are still found pervasively in the environment. Accordingly, the sediments near the outfall record PBDEs and PCBs differently. Accounting for sediment accumulation and mixing, the best fit model of PCB deposition (see Section 2) requires a sudden decrease in the concentration of PCBs in the particles falling onto that site in 1989, the year after the deep water outfall was opened (Fig. 8). PCBs, which are ubiquitously associated with organic matter, are discharged with effluent from the outfall. Concentrations of PCBs in effluent are set by the present PCB background in material entering the treatment plant, and their concentration in the local effluent (9.3-17.4 ng/L, Metro Vancouver, in prep.) is lower than that of PBDEs.

Accordingly, the increased load of particles from effluent discharged near GVRD-3 appears to have diluted the flux of PCBs while at the same time increasing the load of PBDEs.

The core collected at the entrance to Vancouver Harbour (GVRD-2) also shows significant contamination by PCBs and PBDEs (Fig. 6, Table 2). The concentration of PCBs is higher there than at the outfall station (GVRD-3), while that of PBDEs is lower.

There probably is a source of PBDEs bordering the harbour. This sediment core study did not extend into the harbour, but independent measurements show a surface sediment PBDE concentration there similar to that measured near the outfall (Peter Ross, unpublished data, 2007.) Since the sediment accumulation rate is much lower inside the harbour (0.19 g cm-2 yr-1, [START_REF] Johannessen | A sediment and organic carbon budget for the greater Strait of Georgia[END_REF] than near the Iona Island outfall (1.3 g cm -2 yr -1 ), the flux of PBDEs must be lower there too.

Regardless of the actual fluxes to the sediments, benthic organisms are exposed to contaminants, including PCBs, in proportion to concentration (e.g. [START_REF] Reynoldson | Interactions between sediment contaminants and benthic organisms[END_REF].

Consequently, some of the areas of the Strait with the lowest contaminant fluxes may actually be the major points of contaminant entry into the food chain from the sediments. This would be particularly true of PCBs, which are in high concentration where the flux is low, for example in the northern Strait. The highest flux and concentration of PBDEs currently occur together, although that likely will change in the future, after PBDE use is curtailed, and environmental cycling begins to dominate the transport processes as has occurred with PCBs. Even the highest concentration of PCBs measured in these sediments is well below that given as the Interim Sediment Quality Guideline (maximum 2910 pg/g outside Vancouver Harbour vs. guideline 21,500 pg/g, CCME, 2001), and the PCB concentration will continue to decrease, in the absence of a new source. There are no sediment quality guidelines for PBDEs, but, for comparison, the concentration of PBDEs in sediment near the Iona Island outfall (12,800 pg/g) is already more than half the PCB guideline level and increasing rapidly.

So far we have only discussed Total PCBs and Total PBDEs, but each class of chemical is made up of 209 possible congeners. PCBs and PBDEs can be considered in homologue groups, that is, groups of congeners with the same number of chlorine or bromine atoms. The PCB homologues are distributed with a peak in the mid-weight range (Fig. 9a), reflecting the proportions in which the various congeners were used and the number of congeners in each homologue group. This pattern is similar to the distribution found in other environments and organisms (e.g. [START_REF] Ross | High PCB concentrations in free-ranging Pacific killer whales, Orcinus orca: Effects of age, sex and dietary preference[END_REF]. The PBDEs, however, while theoretically comprising the same number of congeners per homologue as the PCBs, are dominated by deca-BDE (Fig. 9b). This formulation, composed principally of BDE-209, is the only form still on the market in North America.

It has an extremely strong affinity for particles (e.g. [START_REF] Law | Levels and trends of brominated flame retardants in the European environment[END_REF]. The next most highly concentrated congeners measured in this study are BDE-47, 99 and 100. These four congeners are predominant in sediments almost everywhere that PBDEs have been measured [START_REF] De Wit | An overview of brominated flame retardants in the environment[END_REF][START_REF] Eljarrat | Occurrence of polybrominated diphenylethers, polychlorinated dibenzo-p-dioxins, dibenzofurans and biphenyls in coastal sediments from Spain[END_REF]. BDE-47, 99 and 100, (but not BDE-209, probably because it is less easily bioaccumulated) also predominate in the biota [START_REF] De Wit | An overview of brominated flame retardants in the environment[END_REF]. These congeners are usually interpreted to represent a mixture of the deca-and penta-formulations. However, one exception is in the coastal sediments off Kuwait, where high atmospheric temperatures (annual range 10-50˚C over land) cause most of the lighter congeners to remain gaseous, and the sediment pattern is dominated by BDE-183, an indicator of the octa formulation that is in use in that country [START_REF] Gevao | Spatial distribution of polybrominated diphenyl ethers in coastal marine sediments receiving industrial and municipal effluents in Kuwait[END_REF] There has been increasing interest in the debromination of deca-BDE (mainly to lighter, more toxic forms, in the environment. As de [START_REF] De Boer | Polybrominated diphenyl ethers in influents, suspended particulate matter, sediments, sewage treatment plant and effluents and biota from the Netherlands[END_REF] point out, debromination of even a very small proportion of deca-BDE in the marine environment would result in significant contamination by the more toxic congeners, because of the high flux of deca-BDE. In laboratory experiments, deca-BDE has been shown to debrominate on exposure to ultraviolet radiation [START_REF] Eriksson | Photochemical decomposition of 15 polybrominated diphenyl ether congeners in methanol / water[END_REF] or zerovalent iron [START_REF] Keum | Reductive debromination of polybrominated diphenyl ethers by zerovalent iron[END_REF]. Debromination is stepwise, one Br substitution at a time, and products include all homologue groups from nona-BDE to mono-BDE. In laboratory conditions the debromination occurs on the time scale of minutes to weeks, depending on the product measured and on the medium in which the deca-BDE is dissolved [START_REF] Eriksson | Photochemical decomposition of 15 polybrominated diphenyl ether congeners in methanol / water[END_REF][START_REF] Keum | Reductive debromination of polybrominated diphenyl ethers by zerovalent iron[END_REF]. Environmental rates of debromination and the distribution of products is unknown, so we did not know which homologues to expect to find in this environment as the end products of debromination. Consequently, we considered changes in the ratios of deca-BDE to all homologue groups as possible indicators of debromination.

To compare congener ratios down-core in all the cores, with their different sedimentation rates, we calculated a time scale for each core as the depth of each measurement divided by that core's sedimentation rate. This quotient does not represent exact years before present because of the extensive benthic mixing of the surface sediment, but it does allow a comparison among cores over approximately equivalent time scales. On this common time scale we plotted the ratio of the sum of the five most abundant congeners (BDE-209, -47, -49, -99, -100) to BDE-209 for each sediment core (Fig. 10a). If debromination were occurring in the sediment, we would expect to see an increase in the BDE ratio downcore, as the lighter congeners were formed from the heavy ones. Confounding this prediction, we would expect the same trend as a result of the pattern of use over time, as the commercial deca formulation has become more dominant. However, Fig. 10 does not show that pattern. An average of the two profiles that extend the furthest back in time shows at best a flat profile. If all the cores were included, the trend would be opposite to that expected for both debromination and changes in usage.

Because we can interpret the sediment cores on a time scale of years to decades, while laboratory results suggest a much shorter half life for deca-BDE, it is possible that the debromination occurs in the environment before the BDEs reach the sediments. We compared the ratios of individual homologues to deca-BDE in effluent from the Iona Island wastewater treatment plant [START_REF] De Wit | An overview of brominated flame retardants in the environment[END_REF]-2005, Metro Vancouver, in prep.) with those in the sediment traps moored nearby (2003)(2004); see Section 2) and those in the sediment cores (collected in 2003). (The mono-, hepta-and octa-BDE homologue groups were frequently undetectable in our samples and were excluded from this calculation.) If debromination had occurred between the effluent and the traps or between the traps (in the water column) and the bottom sediment, we would expect to see the homologue ratios increasing from effluent to traps to surface sediment to sediment averaged over the depth range for which BDE-209 was detectable. Such a pattern in not discernable (Fig. 10b), except possibly for the ratio of nona-to deca-BDE. The apparent increase in the nona-to deca-BDE ratio between the effluent and the sediments might indicate debromination that occurs more slowly in the environment than in the laboratory because of the low intensity of light where the wastewater is discharged (80 m water depth) and the absence of strong reducing agents in the water column. However, it is not clear whether the increase is environmentally significant.

The ratio of the sum of PBDEs to BDE 209 in surface sediments is lowest near the wastewater outfall (GVRD-3) and in the remote northern Strait (GVRD-1), showing greater dominance by BDE-209 at those sites. The higher proportion of BDE-209 near the outfall is not surprising, because that congener's affinity for particles makes it likely to drop out preferentially near the source. If all the other sites had proportionately less BDE-209, the distribution could even be evidence of environmental debromination during the movement through the Strait. However, the high proportion of BDE-209 in the northern Strait is puzzling. Possibly BDE-209 is transported on dust and does not sink until it gets packaged, whereas the lighter BDEs are partitioned into larger, organicrich particles like zooplankton. This would imply that the dust-bound material could travel farther in water than the larger organic particles. A comparison of congener ratios among a larger number of cores might help to resolve this question.

Conclusion

PCBs and PBDEs both are persistent, have or have had local and long-range anthropogenic sources, and have entered a geochemically identical sedimentary regime in the Strait of Georgia. Nevertheless, their distribution in the sediments of the Strait of Georgia differs due to the differing stages in their emission histories. From 1930 until the early 1980s, PCBs were loaded into environmental reservoirs where they have been sorted and redistributed by secondary environmental processes, including volatilization and organic carbon cycling. The amounts cycling are gradually reducing as burial, metabolism and degradation occur, but their cycle is presently controlled by exchanges in which loss from one environmental compartment is accompanied by gain in another. The current distribution of PCBs in Strait of Georgia sediments is largely controlled by sediment flux, bio-mixing and organic/inorganic geochemistry. PCB concentration is high where the flux of diluting, inorganic sediment is low; where the depth of benthic mixing is high, bringing older, more highly contaminated sediment to the surface; and where the concentration of organic carbon, for which PCBs have an affinity, is high.

In contrast, PBDEs are relatively new to the system and growing in importance. Their distribution is strongly controlled by current sources and use patterns; in particular, the entry of these chemicals into domestic dust through the use of treated fabric supports the efficient transport of these compounds to the coastal ocean by municipal wastewater systems. At present the range of surface concentration is similar for the two contaminants, but with PCBs decreasing and PBDEs increasing we expect the latter to predominate in the next decade.

The PBDEs are dominated by BDE-209, and environmental debromination in sediments is not evident. There may be a slight increase in the proportion of nona-to deca-BDE between the effluent and the sediment cores, but more data are required to demonstrate that trend conclusively. PBDE emissions are repeating the experience with PCBs, such that we are now at the same point reached for PCBs in the late 1960s. PBDE discharge continues to increase, and these compounds continue to load into all compartments of the environment. The experience with PCBs suggests that once the discharge of PBDEs stops, there will be a period of readjustment in the sediments that will change the pattern of surface concentration and exposure of the benthos. Eventually, inorganic sediment will bury the PBDEs, but that will take decades after the end of the discharge. 2. Concentrations of selected components of surface sediments and 10-year inventories of total PCBs and total PBDEs. The inventory does not represent exactly 10 years, because benthic organisms mix the surface sediments. The inventory for each core was calculated by multiplying the sedimentation rate by 10 years and summing the contaminants over the resulting depth. It was not possible to calculate the inventory for station 4, because the sedimentation rate is unknown, and we did not analyze all the depth sections for contaminants.

Station

List of Figures

Fig. 1. Emission histories of PCBs and PBDEs, modified from Li et al., 2006 (North American PBDE emission trend) and [START_REF] Rapaport | Historical atmospheric inputs of high molecular weight chlorinated hydrocarbons to eastern North America[END_REF] (% of cumulative emissions of PCBs to Eastern North America) The last five years of the PCB trend (dotted) represent an extrapolation of Rapaport's polynomial fit beyond the date of its publication. Fig. 2. Sediment core sampling locations (GVRD-1 to GVRD-7) and locations of sediment trap moorings (GVRD-A1, GVRD-B1) in the Strait of Georgia. The black line immediately to the east of station GVRD-3 represents the location of the Iona Island outfall pipe. Fig. 3. Profiles of the natural log of the activity of excess 210 Pb in the sediment cores. Dots represent data, solid lines, the accumulation and mixing model, and dotted, horizontal lines, the bottom of the surface mixed layer for each core. Fig. 4. Profiles of the concentration of (a) total PCB and (b) total PBDE in all sediment cores. Fig. 5. Profiles of the concentration of (a) total PCB and (b) total PBDE in sediment core GVRD-1. This core was chosen, because its low sediment accumulation rate means that the entire history of PCB contamination is visible in the 50 cm core. All the cores show the same trends, although they do not all represent the same amount of time. (See Table 1 for sediment accumulation rates.) Error bars are ± 20%, the measurement uncertainty reported by the analytical laboratory. Fig. 6. Distribution of total PCBs and total PBDEs in Strait of Georgia sediments: (a) surface concentration, (b) 10-year inventory, as described in text. Fig. 7. Relationship between the 10-year inventories of (a) total PCB and (b) total PBDE vs. sediment accumulation rate. In panel b, the value for station GVRD-3, near the Iona Island wastewater outfall, is not included in the regression. Fig. 8. Profile of the concentration of Total PCB in sediment core GVRD-3, collected near the Iona Island outfall, with the model fit shown as the solid line. The model only fits the data with a sudden decrease in the PCB concentration of the incident particles in 1989. Error bars are ± 20%, the measurement uncertainty reported by the analytical laboratory. Fig. 9. Homologue distribution in surface sediments averaged over all 7 stations. Because the range of concentrations was so wide, each station's values were meanweighted over all homologues, to preserve the shape of the distribution while giving equal weight to all stations. Fig. 10. (a) Ratio of the sum of the five most abundant congeners 47,49,99,100) to BDE-209 in sediment cores. The horizontal axis represents a time scale calculated as the measurement depth divided by the sedimentation velocity. This does not represent actual years before present, because of benthic mixing, but it is an approximately common time scale for all stations. The heavy black line represents the average of the BDE ratios measured in the two cores that represent the longest time (GVRD-1 and GVRD-2). On this plot, debromination would appear as an increase in the BDE ratio with increasing depth. It is not evident in these data. (b) Homologue ratios in wastewater treatment plant effluent, sediment trap particulates and sediment cores.

ACCEPTED MANUSCRIPT

  concentration of PCBs is highest in the northern Strait (core GVRD-1), where there is no obvious local source, and near the mouth of Vancouver Harbour (GVRD-2). It is low in the central Strait, near the mouth of the Fraser River. In contrast, the concentration of PBDEs near the Iona Island wastewater outfall (GVRD-3, ~12,700 pg/g) dwarfs that measured anywhere else in the Strait (range 270-1800 pg/g). Clearly, the distributions of the two contaminants are determined by different factors. The 10-year inventory at each site shows a different pattern than does surface concentration (c.f. Figs. 6a and 6b), offering a clue to the cause of the distribution observed in the surface concentration plot. Although the northern Strait (core GVRD-1)

  . The concentration of total PBDE measured in the effluent from the Iona Island wastewater treatment plant in 2002-2005 was 61 -432 ng/L (Metro Vancouver. Municipal Wastewater Effluent Characterization. Metro Vancouver,Burnaby, B. C., in prep.), which is higher than that in local river or sea water (lower Fraser River 0.2 -1.8 ng/L in2004-2006, Environment Canada, 2007, unpublished data, southern Strait of Georgia seawater 0.01 -0.02 ng/L[START_REF] Dangerfield | PCBs and PBDEs in the Georgia Basin water column[END_REF]. PBDEs have been measured in high concentration in house dust in Canada and other countries (e.g.[START_REF] Jones-Otazo | Is house dust the missing exposure pathway for PBDEs? An analysis of the urbanfate and Human Exposure to PBDEs[END_REF][START_REF] Wilford | Polybrominated diphenyl ethers in indoor dust in Ottawa, Canada: implications for sources and exposure[END_REF], and their high concentration in wastewater likely reflects the ubiquitous use of flame retardants on household furniture, electronics and toys.

  Figure

Acknowledgements

The authors thank Dr. Peter Ross for his useful insight into the interpretation presented here and Mary O'Brien and Darren Tuele for technical assistance with the sediment traps and cores. We appreciate the assistance of the officers and crew of the CCGS Vector. Flett Research Ltd., in Winnipeg, Canada measured radioisotopes in the sediment cores, which allowed us to calculate the sedimentation and mixing rates. AXYS Analytical Ltd. in Sidney, Canada measured PCBs and PBDEs in the core and trap samples. Dr. Susan Roe of Environment Canada sent and explained the Sediment Quality Guidelines for PCBs. Funding for this work was provided by a collaborative agreement with Metro Vancouver (formerly the Greater Vancouver Regional District) as part of the Strait of Georgia Ambient Monitoring Program, by Environment Canada, through the Georgia Basin Action Plan, and by Fisheries and Oceans Canada.

 1. Sedimentation parameters in sediment cores determined using Pb-210 and an advective-diffusive mixing model. See Fig. 1 for core collection locations. The sediment accumulation rate in core GVRD-3 appears to have increased significantly about 15 years ago. The sedimentation and mixing rates for core GVRD-4 could not be determined from the Pb-210 data, but the sedimentation rate in that core is clearly higher than in any of the others.


Gunther Rosen
email: gunther.rosen@navy.mil

Ignacio Rivera-Duarte

D Bart Chadwick

Adam Ryan

Robert C Santore

Paul R Paquin

Critical tissue copper residues

Keywords: copper, body residue, Biotic Ligand Model, embryo, larval bioassay, Mytilus galloprovincialis, Strongylocentrotus purpuratus, water quality

come   L'archive ouverte pluridisciplinaire

Introduction

Copper (Cu) is frequently measured at elevated concentrations in coastal environments. Its dissolved bioavailability and toxicity to aquatic organisms, however, is dependent upon its physico-chemical speciation [START_REF] Morel | Principles in Aquatic Chemistry[END_REF][START_REF] Campbell | Chapter 2. Interactions Between Trace Metals and Aquatic Organisms: A Critique of the Free-ion Activity Model[END_REF][START_REF] Allen | The importance of trace-metal speciation to water quality criteria[END_REF], which is determined largely by the presence of different organic and inorganic ligands in a given water body [START_REF] Allen | The importance of trace-metal speciation to water quality criteria[END_REF], as well as biological factors such as size, life stage and whether or not dietary uptake is a potential pathway for exposure. Although there is a preponderance of evidence that free aqueous Cu ion is the most biologically available, and therefore potentially toxic, Cu species available for dissolved uptake [START_REF] Sunda | The relationship between cupric ion activity and the toxicity of copper to phytoplankton[END_REF][START_REF] Zamuda | Bioavailability of dissolved copper to the American oyster Crassostrea virginica. I. Importance of chemical speciation[END_REF], ambient water quality criteria (WQC) are currently expressed as dissolved metal (USEPA, 1995a). In addition, WQC are typically developed from toxicity tests conducted in laboratory water that is generally low in metal binding ligands such as dissolved organic carbon (DOC; [START_REF] Usepa | Guidelines for deriving numerical national water quality criteria for the protection of aquatic organisms and their uses[END_REF]. Therefore, national criteria may be either over-protective or under-protective, depending on sitespecific characteristics. Water effect ratio (WER) studies, which involve concurrent toxicity tests with site and laboratory waters, provide a ratio that can be multiplied by the national criterion in order to make it site-specific (USEPA, 1994;USEPA, 2001). WER studies conducted on coastal water bodies have resulted in increases of national WQC for Cu by factors ranging from ~1.5 (e.g. USEPA, 1994;CH2MHill, 2000;[START_REF] Rosen | Use of laboratory toxicity tests with bivalve and echinoderm embryos to evaluate the bioavailability of copper in San Diego Bay, California, USA[END_REF] to 2.7 (City of San Jose, 1998). Although WER studies have proven useful for setting site-specific WQC, they involve labor intensive and costly toxicity testing and associated chemical analyses in order to characterize spatial and temporal differences in the relative dissolved bioavailability and toxicity of Cu at the site.

In freshwater settings, results of WER studies have been successfully predicted using the Biotic Ligand Model (BLM). This approach calculates the critical accumulation of dissolved metal at the site of toxic action using commonly measured water quality parameters such as temperature, dissolved organic carbon (DOC), pH, alkalinity, salinity, Ca, Mg, Na, K, SO 4 -2 , and Cl - (USEPA, 2007). This approach is also more cost-effective than toxicity testing [START_REF] Santore | Biotic Ligand Model of the acute toxicity of metals. 2. Application to acute copper toxicity in freshwater fish and Daphnia[END_REF][START_REF] Santore | Developing Site-Specific Water Quality Criteria for Metals Using the Biotic Ligand Model[END_REF]. The BLM is based on the prediction of metal accumulation by a biotic ligand (e.g. fish gill) that accumulates metal ions from the exposure water. The amount of bound metal is determined by a competition for metal ions with other aqueous ligands, particularly dissolved organic matter (DOM), and the competition for the biotic ligand between the bioavailable forms of the stressor metal and other cations in solution. The utility of the BLM approach to derive site-specific acute water quality criteria for Cu is demonstrated by its incorporation into an updated freshwater WQC document (USEPA, 2007).

For many aquatic invertebrates, the site of toxic action is unknown, or is not accessible for direct measurement. Critical body residues (CBR), based on whole body concentrations, have been suggested as a more toxicologically relevant means of predicting toxicity as compared to water concentrations for a variety of non-polar organic contaminants and metals [START_REF] Mccarty | Enhancing ecotoxicological modeling and assessment[END_REF]. One of the difficulties in applying the CBR approach to metals, however, involves the potential for differences among organisms to compartmentalize, detoxify, and eliminate the nonessential fraction of accumulated metal [START_REF] Mcgeer | Inverse relationship between bioconcentration factor and exposure concentration for metals: Implications for hazard assessment of metals in the aquatic environment[END_REF][START_REF] Vijver | Internal metal sequestration and its ecotoxicological relevance: a review[END_REF][START_REF] Luoma | Why is metal bioaccumulation so variable? Biodynamics as a unifying concept[END_REF]. Although the actual metabolically available concentration of a metal causing toxicity may be the same or similar among species, whole body CBRs may not accurately reflect toxicity due to the species-specific differences in the ability to detoxify and store accumulated metal.

The purpose of this study was to determine if tissue Cu concentrations could be related to toxic effects in early life stages of select marine invertebrates exposed in the laboratory. 1995a, 2003). Because the life stages used are considered non-feeding [START_REF] Sprung | Physiological energetics of mussel larvae (Mytilus edulis). I. Shell growth and biomass[END_REF][START_REF] Fenaux | Variations in the ingestion rate of algal cells with morphological development of larvae of Paracentrotus lividus (Echinodermata: Echinoidea)[END_REF], concerns about the sometimes overlooked role of dietary uptake in metal accumulation [START_REF] Luoma | Prediction of Metal Toxicity in Nature from Bioassays: Limitations and Research Needs[END_REF][START_REF] Fisher | Toxicology tests with aquatic animals need to consider the trophic transfer of metals[END_REF] were reduced. Cu was spiked into seawater samples collected from different locations and seasons in San Diego Bay, CA, USA to demonstrate the effects of varying water quality characteristics on Cu accumulation and toxicity.

Embryo

Methods

Sample collection and treatment

Natural surficial (~ 1 m depth) seawater was collected from the mouth of San Diego Bay (North) and from the southern interior portion of the bay (South; Figure 1). These two sites are non-toxic to bivalve and echinoderm embryos, yet possess different water quality characteristics (e.g. DOC and dissolved Cu concentration) that have been previously demonstrated to correlate with differing degrees of Cu bioavailability [START_REF] Blake | Spatial and temporal variations in Cu speciation in San Diego Bay[END_REF][START_REF] Rivera-Duarte | Copper toxicity to larval stages of three marine invertebrates and copper complexation capacity in San Diego Bay, California[END_REF][START_REF] Rosen | Use of laboratory toxicity tests with bivalve and echinoderm embryos to evaluate the bioavailability of copper in San Diego Bay, California, USA[END_REF]. Seawater was sampled, and filtered to 0.45 µm in the field, using clean techniques (USEPA, 1995b). Samples were held at 4 °C after collection and used within 24 h of collection.

Embryogenesis Exposures

Exposures were conducted according to a slightly modified version of standard toxicity testing procedures for evaluating embryo-larval development success with echinoderms and bivalves (USEPA, 1995c). Modifications to the standard procedures were as follows.

In order to obtain enough biomass for weight measurements and tissue analyses, experiments were conducted at higher initial embryo concentrations (60 and 40 embryos mL -1 , for mussels and sea urchins, respectively). Additionally, larger water volumes (750 mL in 1L polypropylene beakers) were also used. These test conditions were evaluated in preliminary experiments that showed no adverse effects on controls or changes to Cu

Water quality measurements

Water quality parameters including pH, temperature, dissolved oxygen, and salinity were measured daily from surrogate beakers that also included embryos. Dissolved organic carbon (DOC) was determined on unspiked seawater samples according to U.S. EPA Method 415.1.

Larval toxicity assessment

After the appropriate exposure time, each beaker was gently homogenized by swirling with a disposable plastic pipette, and a 5 ml aliquot removed and preserved in 500 µl of concentrated formaldehyde in glass scintillation vials for microscopic evaluation of larval development success. The percentage of normal larvae was determined by counting the first 100 larvae encountered using an inverted microscope at a magnification of 40x. After the designated exposure period, normal mussel larvae have achieved the prodissoconch I stage, characterized by a straight-hinged D-shaped larval shell. Normal sea urchin larvae have reached the pluteus stage, which is pyramidal in shape with four well developed skeletal rods (Figure 2). Because the final larval weight was required for the tissue calculations, final larval weight was also used as a toxicity endpoint.

Whole body residue determination

After removing the toxicity aliquot, the remainder of each sample (~745 mL) was filtered through an acid-cleaned and pre-weighed 8 µm polycarbonate filter in a class-100 clean room. The filtrate was set aside for dissolved Cu measurements (see Section 2.7), which was used to derive water-based toxicity metrics. The filter (containing the larvae) was rinsed with ~100 mL of 18M /cm water to remove loosely bound Cu [START_REF] Playle | Copper accumulation on gills of fathead minnows: Influence of water hardness, complexation and pH of the gill microenvironment[END_REF], and then placed into an acid-cleaned 2 mL polypropylene centrifuge tube. For tissue try weight determination, the contents of the tube were dried at ~25 °C on a hot plate set inside a class 100 all-polypropylene fume hood, until a constant weight was achieved (~48 h).

After recording the mass, the sample was digested with 100 µL Quartz-still grade HNO 3 (Q-HNO 3 ) overnight. The next day, 1 mL 18M /cm water was added, the tube capped, and the sample set aside for Cu measurements (see Section 2.7).

Soft tissue residue determination

Soft tissue Cu concentrations were estimated in a side experiment conducted concurrently with Experiment #3 using North Bay water samples. Whole body residues were measured as previously described (Section 2.5) for 3 replicates, and a duplicate set of 3 replicates was treated with NaOH to digest the tissues, but leave the calcareous parts (larval shell or skeletal rods) intact. After the initial dry weight determination, 250 µL 1 N NaOH was added to the tube. The tube was capped and contents digested overnight. The contents of the tube (including the filter), were then transferred to a new pre-weighed filter, using ~200 mL 18M /cm water for transfer and rinsing. The new filter was allowed to dry for 48 h on a hot plate (~25 °C) for dry weight determination. The sample was then returned to a clean centrifuge tube and acidified with 100 µL Q-HNO 3 and allowed to digest overnight. The next day, 1 mL 18M /cm water was added to result in a final concentration of 1 N Q-HNO 3 , which was stored for Cu analysis (see Section 2.7). The soft tissue concentration was calculated as described in Section 2.8.

Cu Analysis

Dissolved ambient Cu concentrations were measured following clean techniques. Cu in the samples was preconcentrated following the liquid:liquid carbamates preconcentration procedure of [START_REF] Bruland | Analysis of seawater for dissolved cadmium, copper, and lead: An intercomparison of voltammetric and atomic absorption methods[END_REF]. A recovery of 96% (0.57 µg L -1 out of 0.592 ± 0.055 µg L -1 ) in the preconcentration was measured for the standard reference material (SRM) CASS4 (coastal seawater) from the National Research Council of Canada. Cu concentrations in the filtrate and digested tissues were measured by STGFAA spectroscopy by appropriate dilution in 1N Q-HNO 3 and direct injection. A 100% accuracy with a precision of 10% was measured for SRM 1643d (trace metals in water) of the National Bureau of Standards in the STGFAA analysis.

Data analysis

Both exposure water-and residue-based toxicity metrics were calculated with ToxCalc software (Tidepool Scientific) using measured water and tissue concentrations. Water concentrations, therefore, incorporated both background and added Cu. The noobservable-effect-concentrations and residues (NOEC, NOER) were determined using Soft tissue concentrations were estimated using:

st s s wb wb st M M C M C C * * - = (2)
where, C st = Cu concentration in soft tissue (µg g -1 dw), M wb = mass of whole body (µg dw), C s = Cu concentration in shell (µg g -1 dw), M s = mass of shell (µg dw), and M st = mass of soft tissue (µg dw).

Results

Water quality characteristics

A summary of water quality data for the unspiked (no Cu added) samples is provided in Table 1. Parameters varied minimally, with South samples having slightly higher average pH and salinity values compared to North samples. DOC concentrations averaged 1.82 ± 0.59 and 2.84 ± 0.84 mg L -1 for North and South samples, respectively. Dissolved Cu concentrations in unspiked North and South samples, which served as the controls for the experiments, averaged 0.8 ± 0.2 and 2.5 ± 0.0 µg L -1 , respectively.

Water-based toxicity metrics

Cu additions resulted in toxicity at varying water concentrations (Figure 3). Toxicity metrics based on the exposure water concentrations are provided in North Bay Cu EC50 values for both M. galloprovincialis and S. purpuratus were within our laboratory's control chart limits for reference toxicity tests conducted using filtered, natural coastal seawater similar in characteristics to North Bay. This suggests that the organisms used in this study exhibited typical sensitivity from Cu exposure.

Tissue based toxicity metrics -Whole body residues

Whole body residues in controls (no Cu added) tested in North Bay seawater averaged 4.61 ± 2.01 µg g -1 dw for M. galloprovincialis and were 3.63 µg g -1 dw for the one S.

purpuratus sample (Table 2). Whole body concentrations in South Bay seawater controls were somewhat higher than North Bay for M. galloprovincialis but no different for S.

purpuratus (Table 2).

As the water Cu concentration increased, whole body residues of both M. galloprovincialis and S. purpuratus increased (Figure 3), but differed proportionally among individual experiments and between sampling locations (Figure 7). For a given dissolved Cu concentration, higher tissue concentrations were inversely proportional to DOC concentration (Table 1).

Toxicity metrics based on whole body residue concentrations are shown in Table 2 2).

Soft tissue residues

Comparison of tissue dry weights from NaOH digested and non-NaOH digested samples from the same exposure concentrations indicated that the calcareous structures (Figure 2) Although the calcareous structures are a substantial portion of the whole body weight, Cu residues in these structures were almost negligible, averaging only 7.9 and 3.7% of the total residue measured in the whole body for M. galloprovincialis and S. purpuratus, respectively. This resulted in estimates of soft tissue Cu residues that were higher than whole body residues by a factor of 1.6 and 1.2 for M. galloprovincialis and S. purpuratus, respectively (Table 3). Compared to whole body ER50s, soft body ER50s differed less between the species, dropping from a factor of 2.5 to 1.8.

Discussion

Toxicity expressed using water concentrations

These findings substantiate previous work demonstrating both temporal and spatial variability in potential Cu toxicity in San Diego Bay surface water samples. More specifically, this has been demonstrated for embryos of both M. galloprovincialis and S.

purpuratus, when expressed on the basis of dissolved water concentrations [START_REF] Blake | Spatial and temporal variations in Cu speciation in San Diego Bay[END_REF][START_REF] Rivera-Duarte | Copper toxicity to larval stages of three marine invertebrates and copper complexation capacity in San Diego Bay, California[END_REF][START_REF] Rosen | Use of laboratory toxicity tests with bivalve and echinoderm embryos to evaluate the bioavailability of copper in San Diego Bay, California, USA[END_REF]. In the current experiments, we calculated EC50s that differed by a factor of 2.0 for mussel embryos and a factor of 1.4 for sea urchin embryos.

Final larval tissue weight served as an additional toxicity endpoint for M.

galloprovincialis, proving to be nearly as sensitive as the more traditionally used normal development endpoint. This is not surprising, as inhibited formation of the relatively heavy larval shell (46% of total weight in controls at 48 h of development) at toxic concentrations implies that the larvae would not achieve the straight hinged D-shape indicative of normal development. Such a relationship was not apparent in the S.

purpuratus experiments, likely because the skeletal rods of the pluteus are not as significant a factor in the overall mass. In addition, the skeletal structures could still be present, albeit deformed, in abnormal specimens.

The differences in EC50 values between North and South San Diego Bay have previously been attributed to differing water quality characteristics such as total suspended solids (TSS) and DOC concentrations, which tend to be higher in South Bay [START_REF] Blake | Spatial and temporal variations in Cu speciation in San Diego Bay[END_REF][START_REF] Rivera-Duarte | Copper toxicity to larval stages of three marine invertebrates and copper complexation capacity in San Diego Bay, California[END_REF][START_REF] Rosen | Use of laboratory toxicity tests with bivalve and echinoderm embryos to evaluate the bioavailability of copper in San Diego Bay, California, USA[END_REF]. These parameters affect Cu speciation, specifically the concentration of the free ion that is traditionally considered to be the most bioavailable (on a dissolved basis) and toxic form [START_REF] Morel | Principles in Aquatic Chemistry[END_REF][START_REF] Sunda | The relationship between cupric ion activity and the toxicity of copper to phytoplankton[END_REF][START_REF] Eriksen | Copper speciation and toxicity in Macquarie Harbour, Tasmania: an investigation using a copper ion selective electrode[END_REF]. While TSS was not a factor in this study due to filtration of samples prior to exposure, a positive correlation was observed between DOC and M.

galloprovincialis EC50. This is worthy of note considering the relatively narrow range in DOC concentrations that were measured in San Diego Bay. The relationship between M.

galloprovincialis embryo-larval development EC50s and DOC concentration is quite evident for DOC concentrations ranging from <1 to 12 mg L -1 [START_REF] Arnold | Validation and update of a model used to predict Cu toxicity to the marine bivalve Mytilus sp[END_REF], and it is interesting to see this relationship here among so few samples and at relatively low DOC concentrations.

Whole body residue-effect thresholds

The utility of the critical body residue (CBR) approach for Cu was explored in this study by expressing toxicity in terms of whole body concentrations. Whole body ER50s varied little among experiments, as noted by very low CVs and overlapping confidence intervals, suggesting that whole body residues were a better predictor of toxicity than exposure water concentration. This is not unexpected, as tissue measurements reflect only the fraction of Cu available for uptake while dissolved water concentrations do not. CBRs, therefore, have been suggested to be a more appropriate indicator of adverse effects in aquatic biota than external water concentrations [START_REF] Mccarty | Enhancing ecotoxicological modeling and assessment[END_REF]. Here, the apparent value of the CBR approach for Cu in M. galloprovincialis and S. purpuratus larvae is illustrated by very similar toxicity thresholds; a) among samples differing in water quality characteristics; b) between the two species in this study; and c) for other species and life stages, as described below.

The mean ER50s of 49.2 and 142 µg g -1 dw, for M. galloprovincialis and S. purpuratus, respectively, are similar to literature values reported for similar species and life stages. A whole body Cu ER50 of 114 µg g -1 dw was calculated from data reported by [START_REF] Radenac | Bioaccumulation and toxicity of four dissolved metals in Paracentrotus lividus sea-urchin embryo[END_REF] for comparable exposures with embryos of the European sea urchin (Paracentrotus lividus). This is particularly interesting because embryogenesis in P. lividus (mean EC50 = 62.5 µg dissolved Cu L -1 ; [START_REF] Warnau | Spermiotoxicity and embryotoxicity of heavy metals in the echinoid Paracentrotus lividus[END_REF][START_REF] His | A comparison between oyster (Crassostrea gigas) and sea urchin (Paracentrotus lividus) larval bioassays for toxicological studies[END_REF][START_REF] Fernandez | Combined toxicity of dissolved mercury with copper, lead and cadmium on embryogenesis and early larval growth of the Paracentrotus lividus seaurchin[END_REF][START_REF] Radenac | Bioaccumulation and toxicity of four dissolved metals in Paracentrotus lividus sea-urchin embryo[END_REF][START_REF] Lorenzo | Effect of humic acids on speciation and toxicity of copper to Paracentrotus lividus larvae in seawater[END_REF] appears to be considerably less sensitive to Cu than in S. purpuratus (species mean acute value = 12.8 µg dissolved Cu L -1 ; USEPA,

2003) based on water concentrations.

To the best of our knowledge, this is the first time that CBRs have been determined for bivalve larvae. Geffard et al. (2002a), however, observed no effects in 24 h old Pacific oyster (Crassostrea gigas) D-shaped larvae at whole body residues of ~20 µg g -1 dw in sediment elutriate exposures, which is comparable to our NOERs for M. galloprovincialis (18.4 -24.0 µg g -1 dw; Table 2).

Although ER50s differed to some extent between M. galloprovincialis and S. purpuratus, NOERs were essentially identical, suggesting similarity in the thresholds for effects for these two species. The apparent differences in the sensitivity of this endpoint between the species based on ER50 and EC50s, however, may be associated with differences in the ability to regulate or store Cu [START_REF] Luoma | Why is metal bioaccumulation so variable? Biodynamics as a unifying concept[END_REF]. In this study, lower bioconcentration factors (BCFs) observed at sub-effect concentrations in S. purpuratus (1938 ml g -1 ) suggest a lower rate of uptake and/or more efficient elimination as compared to M. galloprovincialis (4574 ml g -1 ) larvae.

Among the criticisms of whole body CBRs as valuable toxicity metrics for metals are that they do not provide an explicit measure of accumulation at the site of action of toxicity (e.g., a particular organ, a membrane bound ion transport system, or enzyme system)(e.g. [START_REF] Vijver | Internal metal sequestration and its ecotoxicological relevance: a review[END_REF]. Similarly, CBRs do not necessarily differentiate between accumulated metal that is metabolically available and that which has been detoxified following uptake.

Inorganic granules [START_REF] Mcgeer | Inverse relationship between bioconcentration factor and exposure concentration for metals: Implications for hazard assessment of metals in the aquatic environment[END_REF][START_REF] Vijver | Internal metal sequestration and its ecotoxicological relevance: a review[END_REF] and metallothionein-like proteins [START_REF] Widdows | Chapter 8. Mussels and environmental contaminants: Bioaccumulation and physiological aspects[END_REF][START_REF] Roesijadi | Metallothionein mRNA accumulation in early developmental stages of Crassostrea virginica following pre-exposure and challenge with cadmium[END_REF][START_REF] Mcgeer | Inverse relationship between bioconcentration factor and exposure concentration for metals: Implications for hazard assessment of metals in the aquatic environment[END_REF][START_REF] Vijver | Internal metal sequestration and its ecotoxicological relevance: a review[END_REF], for example, allow many aquatic organisms to sequester, store, and detoxify high concentrations of metal without observable effects. Metallothioneins have been measured at elevated levels in mussel [START_REF] Roesijadi | Metallothionein mRNA accumulation in early developmental stages of Crassostrea virginica following pre-exposure and challenge with cadmium[END_REF]Geffard et al., 2002b), oyster (Geffard et al. 2002a;[START_REF] Damiens | Metal bioaccumulation and metallothionein concentrations in larvae of Crassostrea gigas[END_REF] and sea urchin [START_REF] Nemer | Developmental regulation, induction, and embryonic tissue specificity of sea urchin metallothionein gene expression[END_REF] larvae following metal exposure originating with embryos. The repeatability of observed effects in the current study at specific whole body concentrations, however, suggests that whole body concentrations are very good surrogates for what is internally available for toxicity, for these species and endpoints. The reason may be that the amount of intracellular Cu that is complexed by metallothionein is small in comparison to the total amount of Cu that has been accumulated. This could be a reflection of the kinetics of metallothionein induction by metals (e.g., Viarengo et al., 1985 Roesijadi and[START_REF] Roesijadi | Kinetic analysis of Cd-binding to metallothionein and other intracellular ligands in oyster gills[END_REF]) and short duration of the exposures used in our laboratory investigations, as well as the absence of the specific organs (e.g., gill and digestive gland) that are present in adults and particularly well suited for the purpose of MT synthesis and detoxification of metals such as Cu (Viarengo et al., 1980). Alternatively, the amount of Cu that interacts at sensitive intracellular biotic ligands may be a relatively small fraction of the total Cu, but proportional to the total amount of Cu that has been accumulated by the larvae.

Interestingly, the CBRs observed for embryo-larval development of M. galloprovincialis are similar to thresholds reported to impact a number of other toxicity endpoints in adult Mytilus sp. [START_REF] Martin | Schema of lethal action of copper on mussels[END_REF] reported a median lethal residue (LR50) of 59 µg g -1 dw in adult M. edulis. Mean sublethal responses (e.g. growth, scope for growth, filtration, reproduction, condition, and change in bioaccumulation) to several bivalve species from 41 different studies were 48.0 and 21.1 µg g -1 dw for effects and no effects residues [START_REF] Salazar | Chapter 9. Linking Bioaccumulation and Biological Effects to Chemicals in Water and Sediment: A Conceptual Framework for Freshwater Bivalve Ecotoxicology[END_REF]. For Mytilus sp. studies specifically, these values were 82.9 and 24.7 µg g -1 dw, respectively. While there is substantial evidence that Mytilus larvae, juveniles, and adults differ in their sensitivity to Cu based on water concentrations [START_REF] Widdows | Chapter 8. Mussels and environmental contaminants: Bioaccumulation and physiological aspects[END_REF], these differences appear to be minimal when expressed as CBRs.

Therefore, as has been proposed with field transplants of adult mussels [START_REF] Salazar | Chapter 9. Linking Bioaccumulation and Biological Effects to Chemicals in Water and Sediment: A Conceptual Framework for Freshwater Bivalve Ecotoxicology[END_REF], our data suggest that assessment of tissue concentrations in M.

galloprovincialis larvae may be a valuable means of predicting biological impacts at sites of concern. Because exposure in filtered laboratory water isn't necessarily indicative of exposure in the field [START_REF] Luoma | Prediction of Metal Toxicity in Nature from Bioassays: Limitations and Research Needs[END_REF], however, validation of these laboratory data should be pursued through field studies.

Although the life stages used in this study are lecithotrophic, the larvae quickly become planktotrophic upon development to veliger and pluteus stages for M. galloprovincialis and S. purpuratus, respectively, presenting potential consequences for use of these CBRs in field exposures. The laboratory experiments conducted here used filtered (0.45 µm)

seawater, but an increase in mussel larval weight was accompanied by the development of the larval shell late in the 48 h exposure (Figure 6). [START_REF] His | Combined effects of temperature and salinity on fed and starved larvae of the Mediterranean mussel Mytilus galloprovincialis and the Japanese oyster Crassostrea gigas[END_REF] also reported growth of M. galloprovincialis larvae in the absence of feeding on phytoplankton in laboratory experiments, and suggested that the larvae were feeding on bacteria. It has also been shown, however, that bivalve veliger larvae have the capacity to directly assimilate dissolved organic material (DOM) from water, allowing them to live well beyond the expected their endogenous energy supply [START_REF] Manahan | The role of dissolved organic material in the nutrition of pelagic larvae: Amino acid uptake by bivalve veligers[END_REF][START_REF] Moran | Physiological recovery from prolonged 'starvation' in larvae of the Pacific oyster Crassostrea gigas[END_REF][START_REF] Baines | Dissolved organic matter and persistence of the invasive zebra mussel (Dreissena polymorpha) under low food conditions[END_REF]). The role that DOM uptake may play in Cu accumulation for these life stages, therefore, needs to be studied.

Background concentrations below NOER

Background larval residues following exposure to Bay samples in the absence of added Cu (controls) were substantially below both NOERs and ER50s derived in this study, providing an additional line of evidence (in addition to water-based site-specific criteria)

that ambient Cu concentrations in San Diego Bay should not negatively impact mussel or sea urchin populations. More specifically, the mean NOERs (20.6 and 23.4 µg g -1 dw;

Table 2) were 3.5 and 6.4 times higher than mean residues observed for mussels (5.90 µg g -1 dw) and sea urchins (3.63 µg g -1 dw) from the control treatments, respectively. Mean background tissue concentrations for both M. galloprovincialis and S. purpuratus in these experiments closely resemble those reported from other laboratory studies using similar life stages and from natural zooplankton samples collected from various geographic locations. Geffard et al. (2002a) reported background Cu concentrations of 5.11 µg g -1 dw for M. edulis larvae following 48 h exposures in clean filtered seawater from the Bay of Arcachon, France, while concentrations in P. lividus pluteus larvae were 5.9 µg g -1 dw in controls using a synthetic seawater [START_REF] Radenac | Bioaccumulation and toxicity of four dissolved metals in Paracentrotus lividus sea-urchin embryo[END_REF]. Whole body tissue Cu concentrations in M. galloprovincialis or S. purpuratus larvae, therefore, appear to be a valuable indicator of Cu exposure at concentrations well below those required to result in abnormal larval development.

Soft tissue ER50

ER50 estimates for soft tissue residues were somewhat higher than those based on whole body residues for both M. galloprovincialis and S. purpuratus larvae. This is due to the relatively low concentrations measured in the larval shell of M. galloprovincialis or the skeletal rods in S. purpuratus. [START_REF] Ringwood | Short-term accumulation of cadmium by embryos, larvae, and adults of an Hawaiian bivalve, Isognomon californicum[END_REF] also reported very low accumulation of Cd in the larval shells of I. californicum, and suggested that the minimal concentrations that were measured were probably a result of adsorption to the shell as opposed to deposition. Low concentrations in the larval shell in our laboratory experiments contrast with generally high concentrations of metals, compared to soft tissues, observed in the exoskeletons (a location for storage and detoxification of metals) of many invertebrates [START_REF] Horowitz | Trace metal concentrations and partitioning in zooplankton, neuston, and benthos from the south Texas outer continental shelf[END_REF]. This could be due to the very short presence, on the order of hours, that these structures existed prior to sampling, thus minimizing the ability for adsorption from the seawater as well as deposition into these structures.

Expressing ER50s as soft tissue reduced the difference in sensitivity between M. galloprovincialis and S. purpuratus larvae to within a factor of two. This was because of the relatively higher proportion between shell and whole body in the mussel larvae as compared to the skeletal rods and whole body for the sea urchin pluteus. Soft tissue ER50s

for M. galloprovincialis rose more than those of S. purpuratus because of the relatively large contribution of shell in the total mass of the 48 h old M. galloprovincialis larva compared to the skeletal rods in the 96 h old S. purpuratus larva.

Comparison to LA50 for fish

The BLM is based on the notion that mortality, or other adverse effects, occurs when the metal-biotic ligand complex reaches a critical concentration [START_REF] Santore | Biotic Ligand Model of the acute toxicity of metals. 2. Application to acute copper toxicity in freshwater fish and Daphnia[END_REF]. The model predicts LC50s or EC50s using easily obtainable water quality parameters rather than using costly and time-consuming toxicity testing. For fish, the concept of the biotic ligand is generally applied to physiologically active receptor sites at the gill surface that are involved in ionoregulatory processes [START_REF] Ditoro | Biotic Ligand Model of the acute toxicity of metals. 1. Technical basis[END_REF]. The Cu ER50 data described herein for M. galloprovincialis and S. purpuratus larvae are considerably higher than the median lethal accumulation (LA50) values developed by [START_REF] Macrae | The copper binding affinity of rainbow trout (Oncorhynchus mykiss) and brook trout (Salvelinus fontinalis) gills: Implications for assessing bioavailable metal[END_REF] for rainbow trout (0.64 µg g -1 ww, or 10 nmol Cu g -1 ww) which were used for calibration of the freshwater BLM [START_REF] Ditoro | Biotic Ligand Model of the acute toxicity of metals. 1. Technical basis[END_REF]. This apparent discrepancy may be due to a few reasons. First, it is already well known that the bioconcentration of metals can vary greatly among different organisms [START_REF] Luoma | Why is metal bioaccumulation so variable? Biodynamics as a unifying concept[END_REF]. Invertebrate zooplankton are known to accumulate metals to particularly high levels [START_REF] Horowitz | Trace metal concentrations and partitioning in zooplankton, neuston, and benthos from the south Texas outer continental shelf[END_REF][START_REF] Fowler | Trace metal monitoring of pelagic organisms from the open Mediterranean Sea[END_REF]Rainbow and White, 1990). Fish, however, closely regulate internal concentrations, and therefore, typically possess lower concentrations than invertebrate zooplankton populations [START_REF] Horowitz | Trace metal concentrations and partitioning in zooplankton, neuston, and benthos from the south Texas outer continental shelf[END_REF][START_REF] Fowler | Trace metal monitoring of pelagic organisms from the open Mediterranean Sea[END_REF].

Perhaps more importantly, however, is that the critical concentrations reported in this study are based on whole and soft body residues, while the fish LA50 is based on accumulation at the specific site of toxic action. Therefore, until the site of toxic action is determined and measured in M. galloprovincialis and S. purpuratus larvae, meaningful comparisons are difficult. In the interim, it is noteworthy that Cu concentration at the site of toxicity appears to be proportional to whole body concentrations for these specific organisms and endpoints.

Conclusions

Based on results of this study, it is apparent that critical whole body residues are a better predictor of Cu toxicity than water concentrations during embryo-larval development of M.

galloprovincialis and S. purpuratus in laboratory exposures. Whole body concentrations are not always good measures of a toxicologically relevant accumulation because they do not distinguish between accumulated metals that are sequestered and stored from those that are biologically available internally for toxicity. The repeatability among CBRs observed in this study for water samples differing in physicochemical characteristics, however, suggests that for these endpoints, tissue residues may provide data that are useful in predicting toxicity in monitoring and assessment studies. Except for their small size, these species and life stages are advantageous and relevant because they often drive the setting of ambient WQC (particularly M. galloprovincialis), they require short exposure times, and the role of dietary uptake is generally considered negligible. Therefore, these data might be useful in decision-making with respect to the regulation of Cu in saltwater, and should be relevant toward the development of a saltwater BLM, which can ultimately be used to predict site-specific metal toxicity. Because of the differences in characteristics among water bodies, these findings should be validated by comparisons with similar experiments in samples from other water bodies that vary more widely in DOC concentration than San Diego Bay. The direct relevance of these data for assessments of field populations also should be investigated.

  -larval development testing of marine invertebrates often plays a critical role in the development of WQC for Cu in saltwater. Identification of a critical accumulation level, could be used to provide calibration data for development of a BLM for seawater, as well as provide a possible alternative approach for monitoring and assessment studies. The endpoints investigated were embryo-larval development of the Mediterranean mussel (Mytilus galloprovincialis) and purple sea urchin (Strongylocentrotus purpuratus). These endpoints were selected because of their role as drivers of the current saltwater WQC for Cu (3.1 µg dissolved Cu L -1 ; USEPA, 1995a) due to effects at very low concentrations (USEPA

  sensitivity. Typically, seven concentrations of Cu bracketing the expected EC50, were prepared by adding Cu (as CuSO 4 •5H 2 O) to the San Diego Bay water samples. A period of 2 to 3 hours was allowed for equilibration of the added Cu prior to the addition of embryos. A filter blank was also included that consisted of unspiked seawater and no embryos. All concentrations were tested in triplicate. Adult mussels were obtained from Carlsbad Aquafarm (Carlsbad, California, USA), where the mussels were cultured under continuously submerged conditions in an embayment about 20 miles north of San Diego. Sea urchins were field collected subtidally from local field populations by Marinus Scientific (Garden Grove, California, USA). Mussels were induced to spawn in the laboratory by thermal shock, which involved raising the water temperature by ~10 °C, while sea urchin spawning was induced by injection of 0.5 M KCl into the peristomal membrane (USEPA, 1995c). Within 4 h of fertilization, embryos were added to test solutions, which were covered, and held on a 16:8 h light:dark cycle at 15 ºC in temperature controlled light chambers (Percival Scientific, Model 136LL). Because the embryos are lecitrophic (yolk supply provides nourishment), test animals were not fed during the exposures, thereby eliminating the dietary exposure pathway. Mussel and sea urchin embryos were allowed to develop for 48 and 96 hours, respectively, prior to termination of the experiments.

  hypothesis testing following confirmation of normal distribution of data and equal variances on arc-sine square root transformed data. Median effects concentrations and residues (EC50, ER50) were determined using Probit point estimation. Sigmoid, 3 parameter curves were used to fit dose responses with Sigma Plot (version 8.0).

  of M. galloprovincialis (shell) and S. purpuratus (skeletal rods) made up approximately 46 and 21% of the final whole body weight in control treatments. The relative fraction of shell or skeletal rod weight decreased at toxic exposure concentrations, which was expected in the mussels due to the lack of shell development at concentrations above the EC50.

Figure 1 .

 1 Figure 1. Map of San Diego Bay, California and approximate location of two surface water sampling area used in this study. The open circle at the mouth of the bay represents "North" samples and the solid circle at the head of the bay represents "South" samples.

Figure 2 Figure 3 .

 23 Figure 2. a) Photograph from inverted light microscope of 48 h old Mediterranean mussel (Mytilus galloprovincialis) larvae under normal conditions, and b) following 24 h digestion with NaOH; and c) photograph of 96 h old purple sea urchin (Strongylocentrotus purpuratus) pluteus larva under normal conditions, and d) following 24 h digestion with NaOH. Actual sizes are approximately 120 and 200 µm, for individual mussel and sea urchin larvae, respectively.

Figure 4 .

 4 Figure 4. A) Median effect concentrations (EC50) and B) median effect residues (ER50) relative to dissolved organic carbon concentration for Mytilus galloprovincialis larvae following 48 h embryo-larval development exposures to copper. Error bars represent 95% confidence limits.

Figure 5 .Figure 6 .Figure 7 .

 567 Figure 5. Comparison of median effect concentrations (EC50) from copper exposures in Mediterranean mussel (Mytilus galloprovincialis) embryo-larval development tests using both normal larval development and dry tissue weight as toxicity endpoints.

  Table 2 and Figure 4a.

	and S. purpuratus, respectively, while DOC concentrations between the two sample
	locations differed by similar factors of 1.32 and 1.38.
	A strong correlation was observed between M. galloprovincialis EC50 values calculated
	from normal larval development data and EC50 values calculated using the final larval
	weight data (r 2 =0.952). As a toxicity endpoint, M. galloprovincialis larval weight was
	slightly less sensitive than normal larval development, with EC50 values averaging 23.5 ±
	M. galloprovincialis EC50s were positively correlated with DOC concentration (r 2 =
	0.995; p = 0.048, n=3). For experiments 2 and 3, in which North and South samples were
	evaluated concurrently, EC50s differed by factors of 1.47 and 1.44 for M. galloprovincialis

M. galloprovincialis EC50 values varied approximately 2-fold, with the lowest EC50s

being observed for North Bay samples, and the highest for the South Bay sample. A similar relationship was observed for the S. purpuratus experiment, but EC50s varied somewhat less (factor of 1.4). The coefficient of variation (CV) among EC50s was 35.1 and 25.5% for M. galloprovincialis and S. purpuratus, respectively. 11.5% higher than those calculated using normal larval development (Figure

5

). Change in weight of the developing mussel embryos over time from a side experiment are shown in Figure

6

. S. purpuratus larval weight EC50s could not be generated due to the low sensitivity of this endpoint.

Table 1 .

 1 Mean (± 1 S.D.) water quality measurements from controls (e.g. no added copper) for each of three larval toxicity and bioaccumulation experiments conducted with surface water from North and South San Diego Bay, CA.

	D.O. a	Temperature	Salinity	DOC b	Cu c

a D.O. = dissolved oxygen b DOC = dissolved organic carbon c Cu= dissolved copper

Table 2 .

 2 Toxicity metrics, based on exposure water concentrations and whole body residues, from seawater samples from North or South San Diego Bay, CA spiked with copper. Toxicity endpoints were normal embryo-larval development of either the Mediterranean mussel (Mytilus galloprovincialis) or purple sea urchin (Strongylocentrotus purpuratus). Metrics include the noobservable-effect concentration (NOEC) and residue (NOER), and median effects concentration (EC50) and residue (ER50). C.L.=confidence limit. Mean control (no added copper) tissue residues are also shown.

				Water concentration (µg Cu L -1 )	Whole body residue (µg Cu g -1 dw)
	Organism	Sample ID	Expt #	NOEC	EC50 (95% C.L.)	Controls	NOER	ER50 (95% C.L.)
	Mussel	North	1	4.10	6.36 (6.22-6.49)	5.84	18.4	50.3 (48.4-52.2)
		North	2	5.34	8.68 (8.47-8.89)	2.29	24.0	44.0 (36.9-52.7)
		South	2	7.08	12.8 (12.6-13.0)	9.65	19.3	53.4 (50.8-56.0)
	Sea Urchin	North	3	9.1	14.3 (13.8-14.9)	3.68	22.9	131 (108-155)
		South	3	14.1	20.6 (20.3-21.0)	3.57	23.8	153 (115-192)

Table 3 .

 3 Comparison of median effect residue (ER50) based on copper content determined in the whole body, the larval shell, and the soft tissue (* indicates estimate) from short-term exposures of copper in North San Diego Bay seawater to embryos of Mediterranean mussels (Mytilus galloprovincialis) and purple sea urchins (Strongylocentrotus purpuratus), in relation to the median effect concentration (EC50).

				ER50 (µg g -1 dw)
	Organism	EC50 (µg L -1 )	Whole body	Shell/Rods	Soft tissue*
	Mussel	9.46	47.7	2.65	76.9
	Sea urchin	14.2	117	1.30	137

Acknowledgements

This study was funded by the U.S. Department of Defense's Environmental Security Technology Certification Program (ESTCP) under project # ER-0523. The authors thank Jennifer Thompson and Christa Zacharias for assistance in the laboratory. Valuable comments from Mike Salazar and two anonymous reviewers improved the quality of the manuscript and are appreciated.