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Keywords:

  • Anadromous fish;
  • Pacific salmonids;
  • Chinook;
  • Water quality standards;
  • PCBs

Abstract

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. METHODS
  5. RESULTS AND DISCUSSION
  6. SUPPLEMENTAL DATA
  7. Acknowledgements
  8. REFERENCES
  9. Supporting Information

In 2011, as part of an update to its state water quality standards (WQS) for protection of human health, the State of Oregon adopted a fish consumption rate of 175 g/day for freshwater and estuarine finfish and shellfish, including anadromous species. WQS for the protection of human health whose derivation is based in part on anadromous fish, create the expectation that implementation of these WQS will lead to lower contaminant levels in returning adult fish. Whether this expectation can be met is likely a function of where and when such fish are exposed. Various exposure scenarios have been advanced to explain acquisition of bioaccumulative contaminants by Pacific salmonids. This study examined 16 different scenarios with bioenergetics and toxicokinetic models to identify those where WQS might be effective in reducing polychlorinated biphenyls (PCBs)—a representative bioaccumulative contaminant—in returning adult Fall chinook salmon, a representative salmonid. Model estimates of tissue concentrations and body burdens in juveniles and adults were corroborated with observations reported in the literature. Model results suggest that WQS may effect limited (< approximately 2 ×) reductions in PCB levels in adults who were resident in a confined marine water body or who transited a highly contaminated estuary as out-migrating juveniles. In all other scenarios examined, WQS would have little effect on PCB levels in returning adults. Although the results of any modeling study must be interpreted with caution and are not necessarily applicable to all salmonid species, they do suggest that the ability of WQS to meet the expectation of reducing contaminant loadings in anadromous species is limited. Integr Environ Assess Manag 2012; 8: 553–562. © 2012 SETAC


INTRODUCTION

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. METHODS
  5. RESULTS AND DISCUSSION
  6. SUPPLEMENTAL DATA
  7. Acknowledgements
  8. REFERENCES
  9. Supporting Information

In 2011, as part of the update of its state water quality standards (WQS) for protection of human health (USEPA 2000), the State of Oregon adopted a fish consumption rate (FCR) of 175 g/day for freshwater and estuarine finfish and shellfish. This value is the highest among all US states and 10 times higher than the US Environmental Protection Agency's (USEPA) national default FCR of 17.5 g/day for the general population (Matzke and Wigal 2011; USEPA 2000). Fish consumption surveys among 4 Native American tribes in the Columbia River basin demonstrated that they consume fish, primarily anadromous, at higher rates than the general population. When Oregon's WQS for organic chemicals and trace metals were calculated using the national FCR, these criteria likely afforded less protection to such high-end consumers of fish and shellfish. Raising the FCR was assumed to offer added protection to populations, such as Native Americans, that consume greater quantities of fish or shellfish on a regular basis and also to specific subpopulations, such as children and women of childbearing age, who may be more susceptible to any chemical contaminants in fish and shellfish. One aspect of increasing the FCR was deciding whether to include consumption of anadromous fish, such as salmon, in the total ingestion rate. USEPA typically does not include salmon in ingestion rate estimates “…on the assumption that adult salmon spend most of their lives in the open ocean and take up bioaccumulative and persistent contaminants almost exclusively via the food chain in that environment.” (USEPA 2007). Nonetheless, as a matter of policy, these species were ultimately included in the data sets used to derive the Oregon FCR because they are of special interest and concern to Northwest Native American tribes (Matzke and Wigal 2011).

Water quality standards for protection of human health that are more stringent, because they are based in part on anadromous fish consumption data, create the expectation that their implementation will lead to lower contaminant levels in such fish. If exposure occurs in waters within the State's jurisdiction (“waters of the state”), then more stringent WQS generated by a higher FCR may reduce both contaminant loads in anadromous fish and risk to humans from subsequent consumption of these fish. This benefit of lower risk, and thus increased availability for consumption, would partially offset regulatory costs associated with what are significantly more stringent WQS. If, however, anadromous species are primarily contaminated in waters beyond the State's jurisdiction (e.g., in the open ocean), then more stringent WQS may simply impose economic and legal costs on the State's economy without the offsetting benefits of reductions in contaminant loads and associated risk. Thus the decision to include anadromous fish in a FCR calculation should be informed by some knowledge of where and when anadromous fish are most likely to be exposed to, and uptake, the majority of their contaminant burden.

Conceptually, contaminant concentration and body burden are a function of where and for how long the specific life stage of a fish and a contaminant are colocated relative to one another in the environment. Because of their anadromous life history, Pacific salmonids occupy 3 distinct habitat types during their lifetimes, each of which may present a different opportunity for exposure to a contaminant: a) freshwater habitats, where eggs hatch and fry develop, b) estuary habitats, where smolts enter marine waters to feed and reside for some time during migration to c) ocean habitats, where the fish spend the majority of their lives. An exposure scenario is defined by where a fish is in space (exposure location), the time it spends in each location (exposure duration), and the contaminant concentration in prey at that location (exposure concentration). A number of exposure scenarios have been advanced, in both the published literature and in anecdotal accounts, to explain the circumstances under which Pacific Northwest salmonids may acquire a contaminant load. Frequently discussed scenarios have contaminant uptake occurring when: a) juveniles are reared in a hatchery (Johnson et al. 2009), b) juveniles (fry, subyearling, yearling) are out-migrating through fresh or estuarine waters (Johnson, Ylitalo, Sloan, et al. 2007; Johnson, Ylitalo, Arkoosh, et al. 2007), particularly if they transit areas with known contamination (Meador et al. 2010), c) adults are in near-shore marine waters (Missildine et al. 2005; O'Neill and West 2009; O'Neill et al. 1998), d) adults are in the open ocean (Cullon et al. 2009; Ewald et al. 1998; Krümmel et al. 2003, 2005; Rice and Moles 2006), e) adults partake of a final “feeding frenzy” in marine waters just before entering freshwater to spawn (anecdotal), or f) adults migrate upriver to spawn (anecdotal). Note that scenario (f) differs from one where exposure is to a contaminant body burden, acquired elsewhere, that is mobilized during spawning (Debruyn et al. 2004). Because the FCR is only relevant to calculation of WQS for protection of human health, this study focused on where fish could acquire tissue residues that could pose a health risk if consumed by humans. It did not address either the protection of aquatic life or the effect of contaminant burdens on the health of anadromous fish, important issues that have been studied by others (Arkoosh et al. 1998; Spromberg and Meador 2005).

The primary objective of this study was to corroborate model estimates of contaminant tissue concentrations and body burdens for specific exposure scenarios with those observed in returning adult Pacific salmonids. Model corroboration considered both the magnitude and lifetime trajectories of both contaminant concentrations and body burdens. A scenario (or scenarios) corroborated by observations might be one that offers a possible explanation for the genesis of those observations. A secondary objective of this study was to identify exposure scenarios within which implementation of a WQS inclusive of anadromous fish might reasonably be expected to reduce contaminant levels in such fish. Because life histories of these anadromous fish are complex and varied, it did not seem possible to test the absolute plausibility of an exposure scenario for all combinations of salmon species, types, evolutionarily significant units (i.e., a population of organisms that is considered distinct for purposes of conservation), or individuals. It did appear feasible, however, to identify a plausible scenario (or scenarios) based on corroboration between scenario-specific model estimates and observed tissue concentrations and body burdens in a representative species of salmonid. Bioenergetics and toxicokinetic (bioaccumulation) models were used to estimate contaminant concentrations and body burdens at locations (spatial dimension) typically occupied by juvenile and adult life stages (temporal dimension) of an idealized individual salmonid.

METHODS

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. METHODS
  5. RESULTS AND DISCUSSION
  6. SUPPLEMENTAL DATA
  7. Acknowledgements
  8. REFERENCES
  9. Supporting Information

Overview

Figure 1 illustrates the conceptual approach to this study. For each simulated day {d} postemergence, a bioenergetics model was used to estimate the mass of invertebrate and vertebrate prey consumed by a fish on that day, for a total lifetime of 2040 d (≈5.5 y). This is likely an overestimate of lifespan, as most Fall chinook return at 3–4 y of age. The model runs for 2040 d to show the potential trajectory of bioaccumulation should a fish live for its theoretical maximum lifespan. Concurrently, the spatial location {y} of the fish on day {d} was estimated based on an individual's idealized life history. The contaminant concentration in prey was quantified at specific locations based on observed levels. A toxicokinetic (bioaccumulation) model was then used to combine estimates of contaminant concentrations in prey with estimates of prey consumption rates to make an estimate of contaminant levels (as both concentration and body burden) in a fish on day {d} at location {y} (Drouillard et al. 2009). Model estimates of tissue concentrations were then compared with those reported in the literature to assess the explanatory power of various exposure scenarios.

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Figure 1. Conceptual model for conduct of this study.

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Representative salmonid

Pacific salmon have evolved many diverse strategies for juvenile migration, estuarine rearing, and adult migration and spawning (Allen and Hassler 1986; Groot and Margolis 1991; Healy 1991; Quinn 2004). Life histories of anadromous salmonids do, however, have some common traits. Adult fish spawn in freshwater streams, usually in late summer or fall. Their large yolky eggs are buried in the substrate, where embryonic development occurs. Juveniles emerge from the substrate the following spring as fry and are dependent on external food sources on emerging. Species life histories diverge at this point, with some species migrating to the estuary and others delaying their migration for months or years. After passing through the estuary, the fish carry out most of the growth in the ocean, spending, depending on the species and stock, between 1 and 6 years there. Adults then return to their natal streams or lakes to spawn and die shortly thereafter.

Fall chinook salmon (Oncorhynchus tshawytscha) were selected as the representative salmonid species because it is highly valued commercially, likely represents an important exposure pathway in the diet of peoples with subsistence lifestyles and high salmon consumption rates, is spiritually and culturally prized among certain Native American tribes, and is known to accumulate contaminants (Carlson and Hites 2005). With chinook, 2 distinct “types” have evolved. A “stream-type” (or Spring) is found most commonly in headwater streams of large river systems. This type has a longer freshwater residency and carries out extensive offshore migrations in the central North Pacific ocean before returning to its natal streams in the spring or summer months. Juveniles migrate as yearlings after overwintering in the river environment. Stream-type juveniles are much more dependent on freshwater stream ecosystems because of their extended residence in these areas, but they spend little time in estuaries before moving to the ocean. They typically spend their first year at sea in near-shore waters before moving into the Gulf of Alaska and Northern Pacific Ocean for 2–4 years. An “ocean-type” (or Fall) is found commonly in coastal streams. Juveniles typically migrate to sea within the first 3 months of life, but timing of migration is quite variable (Reimers and Loeffel 1967). Some disperse to estuaries as fry immediately after emergence, some spend additional time in freshwater before entering the estuary, and some rapidly transit the estuary after short or long periods of residence in freshwater. Ocean-type spend more time in estuaries as juveniles than any other Pacific salmon but variation in timing and duration of estuarine residence is considerable (Hering 2009). After entering ocean waters, they tend to migrate along the coast, spend their ocean life in coastal waters (≈5–8 km offshore), and return to their natal streams or rivers principally as summer and fall runs. After 1–6 years in marine waters, both types return to their natal waters to spawn, the difference being that the ocean-type spawns almost immediately after reaching their natal stream whereas the stream-type typically spends several months in freshwater before spawning. Chinook from Alaska are almost entirely stream-type, whereas those from northern British Columbia are mixed stream- and ocean-types, and those further south in Puget Sound and Oregon waters are predominantly ocean-type (Healy 1991).

Representative contaminant

Polychlorinated biphenyls (PCBs) were selected as the representative persistent, bioaccumulative, and toxic (PBT) contaminant because their physicochemical behavior is well characterized, they have been detected in fresh and marine waters, in the tissues of various salmonid species, and in the tissues of salmonid invertebrate and vertebrate prey (Carlson and Hites 2005; O'Neill and West 2009). Empirical data on PCB concentrations in fresh and marine waters, in chinook salmon, and in their dietary items at various life stages, were drawn from the literature and are summarized in Table S1 (Supplemental Data). These data, although extensive and of good quality, nonetheless had some limitations. Ready comparisons between studies were challenged by PCB concentrations, particularly total concentrations, being reported as differing summations of various aroclors or congeners. Studies did not always report both wet weight and lipid-normalized concentrations, or data needed to convert from one to the other. Body burden estimates, an important adjunct to concentration measurements, were also rarely reported. There were also few measurements of PCB concentrations in adult fish caught at sea at known locations and apparently none of PCBs levels in the stomach contents of such fish. Many studies did not report ancillary data, such as type of salmon (Fall or Spring), weight, length, age, or lipid content, that would have been useful for comparative purposes.

A metaanalysis of the observations listed in Table S1 (Supplemental Data) suggested groups based on sampling location and tissue concentration. Wild fish in headwater (1st–3rd order) streams, including fish in headwater reaches of Puget Sound rivers and large rivers discharging to the ocean, and hatchery fish formed distinct groups, with wild fish in headwater streams having the lowest reported prey and tissue concentrations of all groups. Three groups were evident for Puget Sound: fish collected in the Sound itself, fish collected from contaminated estuaries, and fish taken from presumably un- or less contaminated, estuaries. Fish entering the ocean directly (i.e., not through Puget Sound) from large rivers and those that entered directly from small rivers formed 2 additional groups. Fish caught in the open ocean (e.g., Gulf of Alaska) or in coastal waters outside of Puget Sound (e.g., Johnstone Strait) formed a final group.

From the perspective of observed mean tissue concentrations in returning adults, fish could be placed, seemingly without regard to their type or exposure experience as out-migrating juveniles, into 1 of 3 concentration ranges: 1) a higher “Sound” range (mean tissue concentrations from 35 to 90 µg/kg, w/w) for returning adults whose natal river discharged into Puget Sound, 2) a middle “Large River” range (mean tissue concentrations from 10 to 40 µg/kg, w/w) for returning adults whose natal river discharged directly to the ocean from a watershed with significant urban land use and other anthropogenic impacts (e.g., Columbia River, Fraser River), or 3) a lower “Ocean” range (mean tissue concentrations from 10 to 20 µg/kg, w/w) for returning adults whose natal river also discharged directly to the ocean but from a watershed with few anthropogenic impacts (e.g., Salmon River, OR). This low range also included adults caught in the open waters of the North Pacific, Gulf of Alaska, or Bering Sea on the assumption that their natal rivers were also lightly impacted. Data for returning adults were sufficient to identify 2 ranges based on observed mean body burdens: a high range (200–400 µg/fish, w/w) for adults caught within Puget Sound and in the Duwamish, Deschutes, and Lower Fraser Rivers and a low range (30–100 µg/fish, w/w) for fish caught in the North Pacific Ocean, Gulf of Alaska, or Bering Sea. The differences between these high and low ranges for concentration (6–7-fold) and burden (7–10-fold) may be explained in part by the hydrology of Puget Sound, which is a deep, fjord-like estuary with a narrow connection to oceanic waters through the Strait of Juan de Fuca and shallow sills at Admiralty Inlet. These hydrological features tend to isolate its waters from less contaminated open ocean waters, reduce summer flushing time relative to that of the Strait of Georgia, and allow for contaminants to become entrained within it (Friebertshauser and Duxbury 1972; O'Neill and West 2009; Thomson 1994). This hydrology, combined with considerable urbanization on its surrounding lands, and the presence of several federal Superfund sites, may make the Sound a unique upper bound case for PCB contamination in Pacific Northwest coastal waters. These 5 ranges were used for corroboration purposes, in that a potentially explanatory exposure scenario would be one that placed its model estimates for both tissue concentration and body burden within either the higher or lower ranges observed in returning adult salmon.

Exposure Scenarios

Locations

In general, anadromous fish may be exposed to a contaminant while out-migrating as a juveniles through freshwater or estuarine environments, as adults in the marine environment, or in all 3 environments at different times. Within this general context, measurements summarized in Table S1 (Supplemental Data) were used to identify 16 specific exposure scenarios (Table 1). Because of the known ubiquity of PCBs in aquatic environments, a constant dissolved phased PCB concentration of 10 pg/L in both fresh and marine waters was assumed for all scenarios (Iwata et al. 1993). Seven scenarios (Scenarios 1–8) assumed that exposure occurred via water and prey consumption in only 1 specific location. Scenario 1 had exposure occurring only when wild juveniles out-migrate through river reaches with few, if any, significant anthropogenic impacts (e.g., Salmon River, OR), whereas Scenario 2 assumed exposure only when fish were reared on contaminated food in hatcheries. Exposures in estuaries within Puget Sound could occur when transiting (Scenario 3) a contaminated estuary (e.g., Duwamish Waterway) or another estuary that connects with the Sound (Scenario 4). Estuaries that enter open marine waters directly could be those for large rivers (Scenario 5) with urbanization in their watersheds (e.g., Columbia River, Fraser River), or small rivers (Scenario 6) with little urbanization in their watersheds (e.g., coastal rivers in Washington, Oregon, or Alaska). Exposures could also take place only in Puget Sound (Scenario 7) or only in unconfined coastal or the open marine waters (e.g., Gulf of Alaska) (Scenario 8). Although exposure in just 1 location is possible (e.g., only when transiting a contaminated estuary), fish have the potential, particularly with globally ubiquitous contaminants like PCBs, to be exposed in multiple locations. Eight scenarios (Scenarios 9–16) allowed for combined exposures via water and prey in multiple locations (Table 1). For example, Scenario 15 assumes that an out-migrating wild juvenile (Scenario 1) enters the open ocean (+ Scenario 8) through the estuary of a small river (+ Scenario 6), whereas Scenario 10 assumes that a hatchery-raised fish (Scenario 2) takes up residency in Puget Sound (+ Scenario 7) after entering it through an estuary with known contamination (+ Scenario 4). Although there is no empirical evidence to suggest that Pacific salmon indulge in a prespawning “feeding frenzy” (Higgs et al. 1995), the effect of any such behavior was evaluated by assuming that the consumption rate increased by 10 times for 30 days before the start of the spawning migration. Because Pacific salmon cease feeding during the spawning migration (Higgs et al. 1995), the only uptake of a contaminants during this portion of a salmon's life cycle would be from water via the gills.

Table 1. Summary of exposure scenarios, durations, and concentrations in Fall chinook
ScenarioaExposure locationExposure durationcExposure concentration (CD, µg/kg, w/w)b
JuvenilesdAdultsd
1Freshwater: Wild (upstream of most anthropogenic stressors)1305 (5–23)
2Freshwater: Hatchery13012 (10–14)
3Estuary: Contaminated (Puget Sound)50450 (57–760)
4Estuary: Other (Puget Sound)5034 (22–59)
5Estuary: Large river5062 (20–115)
6Estuary: Small river5010
7Ocean: Puget Sound186028e
8Ocean: Open water18606e
9 (1 + 3 + 7)Wild > contaminated > Sound   
10 (2 + 3 + 7)Hatchery > contaminated > Sound 
11 (1 + 4 + 7)Wild > other > SoundFall chinook resident in Puget Sound
12 (2 + 4 + 7)Hatchery > other > Sound 
13 (1 + 5 + 8)Wild > urban > Ocean   
  • a

    All scenarios assume a constant dissolved phase PCB concentration of 10 pg/L.

  • b

    Concentration as grand mean of means in Table 1 (minimum–maximum range of means).

  • c

    Days postemergence.

  • d

    Concentration in stomach contents.

  • e

    PCB concentration in Pacific herring, assuming contaminated herring is 20% of total adult diet.

14 (2 + 5 + 8)Hatchery > urban > Ocean 
15 (1 + 6 + 8)Wild > non-urban >OceanFall chinook outside Puget Sound
16 (2 + 6 + 8)Hatchery > non-urban > Ocean 
Duration

Chinook in Puget Sound and Oregon waters are predominantly Fall or ocean-type (Healy 1991). Fall chinook may spend approximately 60–210 d postemergence in freshwater and approximately 10 and 90 d in an estuary. For this study, an idealized Fall chinook was assumed to have an exposure duration in freshwater for 130 d postemergence (median of the freshwater range), then in an estuary environment for 50 d (median of the estuary range), and the remaining 1860 d in the marine environment (Table 1), for a total lifetime of 2040 d (≈5.5 y). Median values were selected to explore what happens to a “typical” individual. Here the marine environment was either Puget Sound or the open ocean. A lifetime of this length is expected to overestimate exposure, as returns typically occur within 3–4 years of entering the marine environment.

PCB concentrations in prey

A key input to the toxicokinetic model is the PCB concentration in salmonid invertebrate and vertebrate prey (CD), which is typically the PCB concentration in salmonid stomach contents. Each individual exposure scenario was assigned a different representative value for CD, based on available dietary data as detailed in Table S1 (Supplemental Data) and summarized in Table 1. Its value in all scenarios was a point estimate representing a grand mean. For wild juveniles (Scenario 1), CD was that for fish from the Salmon River (Oregon). For hatchery juveniles (Scenario 2), CD was the mean PCB concentration in feed from various Oregon, Washington, Columbia River, and Columbia Basin fish hatcheries, exclusive of concentrations measured before 2000, as these appeared unusually high relative to more recent measurements. The CD point estimate for a contaminated estuary (Scenario 3) was the mean of data from the Commencement Bay and Duwamish Waterway Superfund sites. That for a large river estuary (Scenario 5) was the mean of data from uncontaminated rivers entering Puget Sound, plus those for the lower Columbia and Fraser Rivers. All of these watersheds include urban lands subject to a variety of anthropogenic stressors, including chemical stressors. The Lower Columbia River, for example, is likely impacted at its confluence with the Willamette River by the Portland metropolitan area (Johnson, Ylitalo, Sloan, et al. 2007; Johnson, Ylitalo, Arkoosh, et al. 2007) and the Fraser River in Canada by its passage through the Vancouver (BC) metropolitan area. For a small river estuary (Scenario 6), CD was the mean in the diet of fish in small, coastal rivers whose estuaries enter the ocean directly, without an intervening sound. Because of a paucity of data on PCB concentrations in adult stomach contents, the CD for adults, in both Puget Sound (Scenario 7) and the open ocean (Scenario 8), was inferred from measured PCB concentrations in Pacific herring (Clupea pallasi), prey comprising 20%–60% of an adult's diet (Healy 1991; West et al. 2008).

Bioenergetic Model (Daily Consumption Rate)

As out-migrating juveniles in freshwater, wild chinook salmon typically feed on pelagic, drifting, and epibenthic larval and adult insects. In estuaries, their diet shifts toward pelagic zooplankton, epibenthic amphipods, and, as they grow larger, small fishes. In the marine environment, the diet of adult chinook is largely comprised of larval and juvenile fishes (principally Pacific herring [Clupea]), pelagic amphipods, and crab megalopa (Healey 1991; Schabetsberger et al. 2003). The Wisconsin bioenergetics model 3.0 (Hanson et al. 1997; Madenjian et al. 2004) was used, unmodified, to estimate the feeding rate (GD), fecal egestion rate (GF), and body weight (W) for juvenile (J) and adult (A) chinook salmon cohorts for each day of a 2040 d lifetime. The model's default values for chinook salmon were used to parameterize its physiological variables (Hanson et al. 1997; Stewart and Ibarra 1991). These included the allometric parameters for dependence of consumption and respiration on body mass, the most sensitive variables (i.e., those with the greatest influence on model predictions). Values for user-specified variables were: weight range (J: 0.1–80 g, A: 80–15 000 g), indigestible fraction of prey (J, A: 20%), prey energy content (J: 2000 J/g, A: 4000 J/g, wet body mass; energy densities of typical prey items (Hanson et al. 1997: see Appendix B), prey dietary fraction (1, unitless; because all prey had the same energy content and digestibility), predator energy content (4000 J/g, wet body mass), and water temperature (10° C, midpoint of optimal growth range [Allen and Haster 1986]).

Toxicokinetic Model (Contaminant Uptake and Retention)

Uptake from prey items and elimination via feces are the major pathways by which fish accumulate and eliminate persistent hydrophobic (log KOW ≈6) organic contaminants such as PCBs (Qiao et al. 2000). Uptake of such contaminants from water via the gills is of less importance due to their generally low concentrations (pg/L) in fresh or marine waters (Gobas and Mackay 1987; Iwata et al. 1993). A mass balance contaminant accumulation model was implemented in STELLA™ (Isee Systems) using variables and algorithms developed by Arnot and Gobas (2003, 2004) and Gobas and Arnot (2010). The 2 most sensitive variables in this model are log KOW (that was a fixed value) and the concentration of a contaminant in prey items (CD). Values for CD (Table 1), as well as the day or days on which a fish is exposed, were varied to match the exposure scenario being evaluated. Table S2 (Supplemental Data) summarizes model variables and equations; relationships for these are shown in Figure S1 (Supplemental Data). This model provided estimates of tissue concentration and body burden resulting for uptake of PCBs from both surface water (via gill exchange) and prey (via consumption) for all, or any portion, of a fish's lifespan. Both estimates were necessary because for non- or poorly metabolized contaminants (such as PCBs) in a fast-growing species, decreases in concentration due mainly to growth dilution may be misinterpreted as reductions in burden (i.e., as a loss of contaminant mass). Burden is a better indicator of the difficult-to-reverse consequences of long term exposure to a recalcitrant contaminant.

RESULTS AND DISCUSSION

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. METHODS
  5. RESULTS AND DISCUSSION
  6. SUPPLEMENTAL DATA
  7. Acknowledgements
  8. REFERENCES
  9. Supporting Information

Scenarios

Observed and model estimated tissue concentrations are listed in Table 2; body burdens in Table 3. Scenario-specific model results, in relation to ranges observed in returning adults after 3–4 years in seawater, are shown in Figure 2 for concentrations and Figure 3 for body burdens. Trajectories through time of tissue concentrations, again in relation to observed ranges, are shown in Figures S2–S8 (Supplemental Material).

Table 2. Comparison of observed and modeled tissue concentrations (µg/kg, w/w)a
ScenarioOut-migrating juvenilesReturning adults
River at 130 dEstuary at 180 d3–4 yb
  • a

    Single and upper values are model estimates; lower values are observed concentrations.

  • b

    Three to four winters in seawater, as average of postemergence Days 1145–1510.

  • c

    Observed tissue concentration, grand mean (range of means).

  • d

    Observed tissue concentration, grand mean (range of means), age of fish not specified.

18 5 (4–8)c62
218 24 (10–50)c153
30.4196 197 (24–725)c38 49 (35–57)d
40.415 45 (40–50)c3 51 (37–83)d
50.427 57 (49–70)c6 34 (11–47)d
60.48 17 (4–46)2 12 (7–19)
70.40.465 67 (40–86)d
80.40.414 11 (9–14)d
98 5 (4–8)c202 197 (24–725)c103 67 (40–86)d
1018 24 (10–50)c211 197 (24–725)c104 67 (40–86)d
118 5 (4–8)c21 45 (40–50)c69 67 (40–86)d
1218 24 (10–50)c30 45 (40–50)c70 67 (40–86)d
136 5 (4–8)c32 57 (49–70)c20 11 (9–14)d
1418 24 (10–50)c42 57 (49–70)c22 11 (9–14)d
158 5 (4–8)c11 17 (4–46)c16 11 (9–14)d
1618 24 (10–50)c19 17 (4–46)c18 11 (9–14)d
Table 3. Comparison of observed and model estimated body burdens (µg/fish, w/w)
ScenarioOut-migrating juvenilesReturning adults
River at 130 dEstuary at 180 d3–4 yk
  • a

    Estimated value in 10 g out-migrating smolts (O'Neill and West 2009).

  • b

    Puget Sound adult chinook after 1–2 winters in saltwater (O'Neill and West 2009).

  • c

    Puget Sound adult chinook after 3–4 winters in saltwater (O'Neill and West 2009).

  • d

    Out-migrating smolts in the Duwamish River, mean (95th percentile) (O'Neill and West 2009).

  • e

    Adults returning to the Duwamish River, mean (95th percentile) (O'Neill and West 2009).

  • f

    Out-migrating hatchery fish in 1989, 1993, and 2000, mean (1 SD) (Meador et al. 2002).

  • g

    Out-migrating wild juveniles collected in 2000 in the Duwamish River upstream of major urban impacts, mean (Meador et al. 2002).

  • h

    Out-migrating juveniles collected in 1989, 1993, 2000 in the Duwamish River estuary, mean (1 SD) (Meador et al. 2002).

  • i

    Adults returning to the Lower Fraser River (BC) and Duwamish and Deschutes Rivers (WA) (Cullon et al. 2009).

  • j

    Adults collected in the Gulf of Alaska and North Pacific Ocean (Carlson and Hites 2005; Easton et al. 2002), in Johnstone Strait (Cullon et al. 2009), and off Vancouver Island (BC) (Ikonomou et al. 2007).

  • k

    3 to 4 winters in seawater, as average of post-emergence days 1145 to 1510.

10.20.49
 0.2g0.4a 
211 2.0 (1.5)f18
30.0112212
  2.1 (9.2)d350 (800)e
  4.8 (0.8)h218–333i
40.01119
50.01232
60.010.511
70.010.02372
   260–340b
   280–390c
80.010.0282
   29–98j
90.212587
100.512596
110.21394
120.52403
130.22116
140.52126
150.2193
1611102
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Figure 2. Comparison of tissue concentration estimates (µg/kg, w/w) by scenario (vertical bars) to observed tissue concentration ranges (boxes with dotted line borders) in adult fish at 3–4 y in marine waters.

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Figure 3. Comparison of body burden estimates (µg/fish) by scenario (vertical bars) to observed body burden concentration ranges (boxes with dotted line borders) in adult fish at 3–4 y in marine waters.

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Individual scenarios

Exposures only in upstream freshwater habitats (Scenario 1 [Figure S2]) or only in a hatchery (Scenario 2 [Figure S2]), or only while transiting uncontaminated estuaries (Scenarios 4 and 6 [Figure S3]) all failed to produce estimates for returning adult fish within any of the observed concentration ranges. With the unusually high dietary concentrations reported before 1993 set aside, hatchery fish exposed to contaminated food (Scenario 2) would be indistinguishable from those exposed only in the open ocean (Scenario 8 [Figure S4]), which suggests that hatcheries may be an unlikely sole source of PCB loads in returning adults. Exposure only in Puget Sound (Scenario 7 [Figure S4]) or only in the open ocean (Scenario 8) was sufficient to generate concentrations and burdens within the Sound and Ocean ranges, respectively, for fish with 3–4 years in seawater and also at end-of-life (Figures 2 and 3). Exposure only in a large river estuary (Scenario 5 [Figure S3]) yielded concentration estimates just below, but burden estimates at, the lower range, whereas short, intense exposures not unlike those achieved by passage through an estuary containing an in-water contaminated site (Scenario 3 [Figure S3]) generated concentration and burden estimates at the higher range. An estuary fed by a large river flowing through areas impacted by anthropogenic chemical stressors (e.g., the Columbia River at its confluence with the Willamette River is affected by the urban areas of Portland [OR] and Vancouver [WA], as well as a large in-water Superfund site) could produce such exposures.

Scenario 3, a single large exposure when out-migrant juveniles transit a contaminated estuary, produced tissue and burden estimates at the lower bound of those observed in 3–4-year-old adults (Figures 2 and 3). Tissue concentrations slowly declined but body burdens were recalcitrant, indicating that even a relatively brief (≤50 d) exposure to elevated prey concentrations may have lasting consequences in terms of increased PCB burdens carried by adults. Average and adjusted average contaminant concentrations in adults taken from Puget Sound after 1–4 years in salt water (O'Neill and West 2009) were compared to model estimated concentrations. Scenario 3 provided the closest approximation to these observations in terms of both magnitude of, and rate of decline in, concentration (Figure S9). Colloquially, a juvenile transiting a contaminated location appears to “jump-start” acquisition of a body burden of highly bioaccumulative contaminants such as PCBs. Scenario 7 did not suggest declines in tissue concentrations, suggesting that apparent declines, particularly in burdens, result from samples composed of individual fish with differing exposure experiences.

Absent a short, intense exposure to chemical stressors (e.g., Scenarios 3 and 5), simple residence in, or extended transit through, contaminated marine waters may be sufficient to generate the majority of the observed loads. The degree of loading may be a function of the extent of the time spent is residence or transit. Thus exposures before entering marine waters that do not involve intense exposures are unlikely to be the principal source of PCB loads observed in returning adults. This finding of a dominant role for exposure in open marine waters is consistent with reports by others (O'Neill and West 2009) and with the USEPA rationale for not including anadromous fish in exposure estimates (USEPA 2007).

Multiple scenarios

Exposure upstream, then in a contaminated estuary, then in Puget Sound (Scenarios 9 and 10 [Figure S5]), approximately doubled concentration and burden estimates over those for the Sound (Scenario 7) alone. Conversely, exposures upstream, then in an uncontaminated Puget Sound estuary, then in Puget Sound (Scenarios 11 and 12 [Figure S6]) did not produce concentration and burden estimates different than those in the Sound (Scenario 7) alone. This emphasizes the role played by short but intense exposures associated with a contaminated estuary. Exposure upstream, then in an large river estuary, then in open marine waters (Scenarios 13 and 14 [Figure S7]), also approximately doubled concentration and burden estimates over those for the open ocean (Scenario 8) alone. An initial spike in concentration (Figure S7) due to the large river estuary was subsequently ameliorated by a longer exposure to less contaminated prey in the open ocean, causing concentration and burden estimates to settle into the large river range between the Sound and Ocean ranges. However, exposures upstream, then in a small river estuary, then in the open ocean (Scenarios 15 and 16 [Figure S8]) did not produce concentration and burden estimates different than those in the open ocean (Scenario 8) alone.

Miscellaneous scenarios

Uptake from water was a small and comparatively inconsequential source of PCB concentrations and burdens, indicating that observed adult burdens could not be obtained only during the upstream spawning migration in freshwater. However, the ubiquity and persistence of legacy PCBs (and other legacy chemicals with similar physicochemical properties) virtually guarantees that any fish, whether anadromous or resident, will have some small PCB burden (CTUIR 2007; Henny et al. 2003). The “feeding frenzy” scenario (a 10-fold increase in the consumption rate in the 30 d before the start of spawning) had no discernible effect on either concentrations or burdens estimated for any scenario. It is thus highly unlikely that an adult fish could acquire concentrations or burdens on the order of observed levels simply by sharply increasing its feeding near the end of its life.

Implications for WQS

This study focused on 1 type (Fall) of 1 species (chinook) of salmonid and on 1 PBT contaminant (PCBs), corroborated by a metaanalysis of available data on tissue concentrations and body burdens of that contaminant collected in the Pacific Northwest and Alaska by a number of researchers. It applied these data to models that are well established and whose behavior, including sensitive inputs, is well understood. However, these species, data, and models all embed uncertainties of various types, not all of which are readily identifiable or quantifiable. As a result, the results of this (or any) modeling study must be interpreted with caution; however, it may still provide insights into the efficacy of WQS for reducing contaminant loads in Fall chinook salmon. Note, however, that the specific applicability of these results to other salmonid species was not determined here. At a minimum, these results may be useful for dispelling assertions about exposure scenarios that are physiologically improbable and whose pursuit is unlikely to result in protective outcomes.

Results suggest that using WQS as waterbody target concentrations may yield only small (≤2 ×) reductions in PCB levels (or of other ubiquitous legacy contaminants with similar PBT properties) in returning adult Fall chinook salmon because the majority of uptake likely occurs while adults are in marine waters beyond the state's jurisdiction. WQS would also have little effect on hatchery fish whose PCB load stems from consumption of contaminated feed (Scenario 2). Scenario 8 results suggest, as have others (O'Neill and West 2009), that PCB loads in fish either resident outside of Puget Sound or not in contact with a contaminated estuary likely stem primarily from exposure in open marine waters. Because states do not have jurisdiction over the open ocean, implementation of WQS will not occur in such waters.

Puget Sound is unique marine water body in that it is both poorly flushed and subject to contaminant loading from surrounding urban landscapes, which have been shown to be disproportionate contributors of chemical stressors (Black et al. 2000; Paul and Meyer 2001). It is also host to several major in-water contaminated sites, where WQS are currently being used to guide remediation efforts. Although addressing these sites is likely to eliminate excesses in concentrations and burdens (e.g., Scenarios 9 and 10), doing so is unlikely to result in large reductions in bioaccumulative contaminants in anadromous fish. Because of the known relationship between urban land use and chemical stressors (Black et al. 2000), use of WQS in controlling or reducing contaminants from nonpoint sources (e.g., runoff from impervious surfaces, nonpermitted stormwater flows, runoff of air deposition [Hope 2008]) will also be required (McCarthy et al. 2008). Because permitted and properly managed point sources (e.g., industrial, wastewater treatment, permitted stormwater) are no longer significant contributors of PCBs to watersheds, use of WQS to regulate such sources would not reduce chemical loadings throughout the Sound. Although implementation of WQS for all waters entering the Sound may, over time, yield lower contaminant levels within the Sound, there are likely to be practical limits on the affect WQS can have on globally distributed legacy contaminants such as PCBs.

Including anadromous fish in the FCR used for developing WQS for protection of human health creates the expectation that implementation of this WQS will significantly reduce bioaccumulative contaminants (e.g., PCBs, PBDEs, dioxins/furans) in such fish. Based on these model results, meeting WQS may lead to small reductions (≤2 ×) only for returning Fall chinook salmon adults that were resident in a confined water body (e.g., Puget Sound) or who transited a highly contaminated estuary (e.g., Duwamish Waterway) as out-migrating juveniles. Otherwise, it may be unrealistic to expect attainment of WQS to result in reduced contaminant burdens in species who receive these burdens as adults in unconfined coastal or open marine waters. Where attainment of WQS can be physically linked to reductions in contaminant loads, benefits will typically out-weigh costs associated with its attainment. Conversely, any physical disconnects between attainment of WQS and expected reductions in contaminant loads creates a situation with costs but few, if any, off-setting benefits. Such a cost-benefit disparity can frustrate those seeking the protection of WQS and those legally required to implement controls designed to attain it.

SUPPLEMENTAL DATA

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. METHODS
  5. RESULTS AND DISCUSSION
  6. SUPPLEMENTAL DATA
  7. Acknowledgements
  8. REFERENCES
  9. Supporting Information

Supporting Tables S1–S9.

Supporting Figure S1.

Supporting Figure S2.

Acknowledgements

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. METHODS
  5. RESULTS AND DISCUSSION
  6. SUPPLEMENTAL DATA
  7. Acknowledgements
  8. REFERENCES
  9. Supporting Information

This manuscript benefited from the constructive comments of 2 anonymous reviewers. All opinions expressed herein are solely those of the author and do not necessarily represent ODEQ policy or guidance, or those of any other public or private entity. No official endorsement is implied or to be inferred.

REFERENCES

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. METHODS
  5. RESULTS AND DISCUSSION
  6. SUPPLEMENTAL DATA
  7. Acknowledgements
  8. REFERENCES
  9. Supporting Information
  • Allen MA, Hassler TJ. 1986. Species profiles: Life histories and environmental requirements of coastal fishes and invertebrates (Pacific Southwest)—Chinook salmon. Biological Report 82 (11.49). Washington (DC): US Fish and Wildlife Service.
  • Arkoosh MR, Casillas E, Clemons E, Kagley AN, Olson R, Reno P, Stein JR. 1998. Effect of pollution on fish diseases: Potential impacts on salmonid populations. J Aq Animal Health 10: 182190.
  • Arnot JA, Gobas FAPC. 2003. A generic QSAR for assessing the bioaccumulation potential of organic chemicals in aquatic food webs. QSAR Comb Sci 22: 337345.
  • Arnot JA, Gobas FAPC. 2004. A food web bioaccumulation model for organic chemicals in aquatic ecosystems. Environ Toxicol Chem 23: 23432355.
  • Black RW, Haggland AL, Voss FD. 2000. Predicting the probability of detecting organochlorine pesticides and polychlorinated biphenyls in stream systems on the basis of land use in the Pacific Northwest, USA. Environ Toxicol Chem 19: 10441054.
  • Carlson DL, Hites RA. 2005. Polychlorinated biphenyls in salmon and salmon feed: Global differences and bioaccumulation. Environ Sci Technol 39: 73897395.
  • [CTUIR] Confederated Tribes of the Umatilla Indian Reservation. 2007. Umatilla River Fish Toxicant Studies. Confederated Tribes of the Umatilla Indian Reservation, Pendleton, OR. Report prepared by Jones and Stokes, Portland, OR.
  • Cullon DL, Yunker MB, Alleyne C, Dangerfield NJ, O'Neill S, Whiticar MJ, Ross PS. 2009. Persistent organic pollutants in chinook salmon (Oncorhynchus tshawytscha): Implications for resident killer whales of British Columbia and adjacent waters. Environ Toxicol Chem 28: 148161.
  • Debruyn AMH, Ikonomou MG, Gobas FAPC. 2004. Magnification and toxicity of PCBs, PCDDs, and PCDFs in upriver-migrating Pacific salmon. Environ Sci Technol 38: 62176224.
  • Drouillard KG, Paterson G, Haffner GD. 2009. A combined food web toxicokinetic and species bioenergetic model for predicting seasonal PCB elimination by yellow perch (Perca flavescens). Environ Sci Technol 43: 28582864.
  • Easton MDL, Luszniak D, Von der Geest E. 2002. Preliminary examination of contaminant loadings in farmed salmon, wild salmon and commercial salmon feed. Chemosphere 46: 10531074.
  • Ewald G, Larsson P, Linge H, Okla L, Szarzi N. 1998. Biotransport of organic pollutants to an inland Alaska lake by migrating Sockeye salmon (Oncorhynchus nerka). Arctic 51: 4047.
  • Friebertshauser MA, Duxbury AC. 1972. A water budget study of Puget Sound and its subregions. Am Soc Limnol Oceanog 17: 237247.
  • Gobas FAPC, Arnot JA. 2010. Food web bioaccumulation model for polychlorinated biphenyls in San Francisco Bay, California, USA. Environ Toxicol Chem 29: 13851395.
  • Gobas FAPC, Mackay D. 1987. Dynamics of hydrophobic organic chemical concentration in Fish. Environ Toxicol Chem 6: 495504.
  • Groot C, Margolis L. 1991. Pacific salmon life histories. Vancouver (BC): University of British Columbia Press. 564 p.
  • Hanson PC, Johnson TB, Schindler DE, Kitchell JF. 1997. Fish Bioenergetics 3.0 for Windows. WISCU-T-97-001. Madison Center for Limnology and Sea Grant Institute, University of Wisconsin, Madison, WI.
  • Healy MC. 1991. Life history of chinook salmon (Oncorhynchus tshawytscha). In: Groot C, Margolis L, editors. Pacific salmon life histories. Vancouver (BC): University of British Columbia Press. p 311391.
  • Henny CJ, Kaiser JL, Grove RA, Bentley VR, Elliott JE. 2003. Biomagnification factors (fish to osprey eggs from Willamette River, Oregon, U.S.A.) for PCDDs, PCDFs, PCBs, and OC pesticides. Environ Monitor Assess 84: 275315.
  • Hering DK. 2009. Growth, residence, and movement of juvenile chinook salmon within restored and reference estuarine marsh channels in Salmon River, Oregon [MS thesis]. Corvallis (OR): Oregon State Univ.
  • Higgs DA, Macdonald JS, Levings CD, Dosannjh BS. 1995. Nutrition and feeding habits in relation to life history stage. In: Groot C, Margolis L, Clarke WC, editors. Physiological ecology of Pacific salmon. Vancouver (BC): University of British Columbia Press. p 174176.
  • Hope BK. 2008. A model for the presence of polychlorinated biphenyls (PCBs) in the Willamette River Basin (Oregon). Environ Sci Technol 42: 59986006.
  • Ikonomou MG, Higgs DA, Gibbs M, Oakes J, Skura B, McKinley S, Balfry SK, Jones S, Withler R, Dubetz C. 2007. Flesh quality of market-size farmed and wild British Columbia Salmon. Environ Sci Technol 41: 437443.
  • Iwata H, Tanabe S, Sakai N, Tatsukawa R. 1993. Distribution of persistent organochlorines in the oceanic air and surface seawater and the role of ocean on their global transport and fate. Environ Sci Technol 27: 10801098.
  • Johnson LL, Willis ML, Olson OP, Pearce RW, Sloan CA, Ylitalo GM. 2009. Contaminant concentrations in juvenile fall Chinook salmon from Columbia River hatcheries. N Am J Aquaculture 72: 7392.
  • Johnson LL, Ylitalo GM, Arkoosh MR, Kagley AN, Stafford C, Bolton JL, Buzitis J, Anulacion BF, Collier TK. 2007. Contaminant exposure in outmigrant juvenile salmon from Pacific Northwest estuaries of the United States. Environ Monitor Assess 124: 167194.
  • Johnson LL, Ylitalo GM, Sloan CA, Anulacion BF, Kagley AN, Arkoosh MR, Lundrigan TA, Larson K, Siipola M, Collier TK. 2007. Persistent organic pollutants in outmigrant chinook salmon from the Lower Columbia Estuary, USA. Sci Tot Environ 374: 342366.
  • Krümmel EM, Gregory-Eaves I, Macdonald RW, Kimpe LE, Demers MJ, Smol JP, Finney B, Blais JM. 2005. Concentrations and fluxes of salmon-derived polychlorinated biphenyls (PCBs) in lake sediments. Environ Sci Technol 39: 70207026.
  • Krümmel EM, Macdonald RW, Kimpe LE, Gregory-Eaves I, Demers MJ, Smol JP, Finney B, Blais JM. 2003. Delivery of pollutants by spawning salmon. Nature 415: 255256.
  • Mandenjian CP, O'Connor DV, Chernyak SM, Rediske RR, O'Keefe JP. 2004. Evaluation of a chinook salmon (Oncorhynchus tshawytscha) bioenergetics model. Can J Fish Aq Sci 61: 627635.
  • Matzke A, Wigal J. 2011. Issue paper: Human Health toxics criteria, human health toxics rulemaking, 2008-2011. Portland (OR): Water Quality Division, Oregon Department of Environmental Quality.
  • McCarthy SG, Incardona JP, Scholz NL. 2008. Coastal storms, toxic runoff, and the sustainable conservation of fish and fisheries. Am Fish Soc Symp 64: 727.
  • Meador JP, Collier TK, Stein JE. 2002. Use of tissue and sediment-based threshold concentrations of polychlorinated biphenyls (PCBs) to protect juvenile salmonids listed under the U.S. Endangered Species Act. Aq Conser Mar Fresh Ecosys 12: 493516.
  • Meador JP, Ylitalo GM, Sommers FC, Boyd DT. 2010. Bioaccumulation of polychlorinated biphenyls in juvenile chinook salmon (Onchorhynchus tshwytscha) outmigrating through a contaminated estuary: dynamics and application. Ecotoxicology 19: 141152.
  • Missildine BR, Peters RJ, Chin-Leo G, Houck D. 2005. Polychlorinated biphenyl concentrations in adult Chinook salmon (Oncorhynchus tshawytscha) returning to coastal and Puget Sound hatcheries of Washington State. Environ Sci Technol 39: 69446951.
  • O'Neill SM, West JE, Hoeman JC. 1998. Spatial trends in the concentration of polychlorinated biphenyls (PCBs) in Chinook (Oncorhynchus tshawytscha) and Coho salmon (O. kisutch) in Puget Sound and factors affecting PCB accumulation: Results from the Puget Sound Ambient Monitoring Program. Olympia (WA): Puget Sound Research Proceedings. p 312328.
  • O'Neill SM, West JE. 2009. Marine distribution, life history traits, and the accumulation of polychlorinated biphenyls in Chinook salmon from Puget Sound, Washington. Trans Am Fish Soc 138: 616632.
  • Paul MJ, Meyer JL. 2001. Streams in the urban landscape. Ann Rev Ecol Syst 32: 333365.
  • Qiao P, Gobas FAPC, Farrell AP. 2000. Relative contributions of aqueous and dietary uptake of hydrophobic organic chemicals to the body burden in juvenile trout. Arch Environ Contam Toxicol 39: 369377.
  • Quinn TP. 2004. The behavior and ecology of Pacific salmon and trout. Seattle (WA): University of Washington Press. 320 p.
  • Reimers PE, Loeffel RE. 1967. The length of residence of juvenile fall chinook salmon in selected Columbia River tributaries. Res Briefs Fish Comm Oregon 13: 519.
  • Rice S, Moles A. 2006. Assessing the potential for remote delivery of persistent organic pollutants to the Kenai River in Alaska. Alaska Fish Res Bull 12: 153157.
  • Schabetsberger R, Morgan CA, Brodeur RD, Potts CL, Peterson WT, Emmett RL. 2003. Prey selectivity and diel feeding chronology of juvenile chinook (Oncorhynchus tshawytscha) and coho (O. kisutch) salmon in the Columbia River plume. Fish Oceanog 12: 523540.
  • Spromberg JA, Meador JP. 2005. Relating results of chronic toxicity responses to population-level effects: Modeling effects on wild chinook salmon populations. Integr Environ Assess Manag 1: 921.
  • Stewart DJ, Ibarra M. 1991. Predation and production by salmonine fishes in Lake Michigan, 1978-88. Can J Fish Aq Sci 48: 909922.
  • Thomson RE. 1994. Physical oceanography of the Strait of Georgia-Puget Sound-Juan de Fuca Strait system. Can Tech Rep Fish Aq Sci 1948: 3648.
  • [USEPA] US Environmental Protection Agency. 2007. Framework for selecting and using tribal fish and shellfish consumption rates for risk-based decision making at CERCLA and RCRA cleanup sites in Puget Sound and the Strait of Georgia (Working Document). Office of Environmental Cleanup, Office of Air, Waste, and Toxics, Office of Environmental Assessment, Seattle (WA): USEPA, Region 10.
  • [USEPA] US Environmental Protection Agency. 2000. Methodology for deriving ambient water quality criteria for the protection of human health. Washington DC: USEPA. EPA-822-B-00-004.

Supporting Information

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. METHODS
  5. RESULTS AND DISCUSSION
  6. SUPPLEMENTAL DATA
  7. Acknowledgements
  8. REFERENCES
  9. Supporting Information

Additional Supporting Information may be found in the online version of this article.

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