Port of Toledo, Oregon

LEVEL II (SCREENING) ECOLOGICAL RISK ASSESSMENT – FRED WAHL BOATYARD

Prepared for: Port of Toledo, Oregon 385 NW First Street, Unit 1 Toledo, Oregon 97391

For submittal to: Oregon Department of Environmental Quality Western Region 165 East 7th Avenue, Suite 100 Eugene, Oregon 97401

July 12, 2010

Prepared by:

200 West Mercer Street, Suite 401  Seattle, Washington  98119

Table of Contents

List of Tables i List of Figures ii Acronyms iii 1 Introduction 1 1.1 SITE HISTORY 4 1.2 REGULATORY STATUS 4 1.3 SUMMARY OF LEVEL I ERA RESULTS 4 1.3.1 Contaminants of Interest 4 1.3.2 Potential Ecological Receptors 4 1.3.3 Primary Exposure Pathways 5

2 Site Survey 6 3 Sampling Results and Screening Results and Analysis 7 3.1 SITE DESCRIPTION 7 3.2 SITE-SPECIFIC ECOLOGICAL RECEPTORS 8 3.2.1 Salmonids 9 3.2.2 Sturgeon 10 3.2.3 Invertivorous fish 10 3.2.4 Benthic invertebrate community 11 3.3 CONTAMINANTS OF POTENTIAL ECOLOGICAL CONCERN 12 3.4 PRELIMINARY CONCEPTUAL SITE MODEL 16 3.4.1 Known ecological effects 16 3.4.2 Candidate assessment endpoints 18 3.4.3 Relevant and complete exposure pathways 18

4 Conclusions and Recommendations 23 5 References 24

Appendix A. ProUCL Output Results

List of Tables

Table 3-1. Aquatic habitats, supported species, and selected ecological receptors 8 Table 3-2. Sediment screening values 13 Table 3-3. Sediment sampling results compared to SLVs 15

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List of Figures

Figure 1-1. Site location 2 Figure 1-2. Survey area 3 Figure 3-1. Conceptual site model 20

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Acronyms

Acronym Definition COI contaminant of interest CPEC contaminant of potential ecological concern EC environmental concentration EF exceedance factor EPA US Environmental Protection Agency ERA ecological risk assessment ESA Endangered Species Act NOAA National Oceanic and Atmospheric Administration ODEQ Oregon Department of Environmental Quality ODFW Oregon Department of Fish and Wildlife SLV screening-level value TBT tributyltin TOC total organic carbon TMDP technical management decision point UCL upper confidence limit USFWS US Fish and Wildlife Service

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1 Introduction

This report presents the findings of a Level II (screening) ecological risk assessment (ERA) conducted at the Yaquina River shoreline along the Fred Wahl Boatyard (Figure 1-1). The Level II ERA was performed for the area delineated by the approximate survey boundary (Figure 1-2), hereafter referred to as the Site. The Level I (scoping) ERA conducted at the boatyard in June 2010 identified ecological receptors that might use the habitats at the Site, and potentially relevant and complete pathways to benthic invertebrates and fish. Based on these findings and in accordance with the Guidance for Ecological Risk Assessment: Levels I, II, III, IV (ODEQ 1998, 2001), a Level II ERA was conducted at the Site. This report is organized according to Oregon Department of Environmental Quality (ODEQ) guidance (1998, 2001). Section 1 includes limited background information to provide context for the Level II ERA. Section 2 describes the results from the Level II survey conducted to characterize the aquatic habitats and ecological receptors of interest to the screening assessment. Section 3 describes the results of previous sediment sampling and screening and analysis of the data relative to ecological factors applicable to the Level II assessment. Conclusions and recommendations are presented in Section 4. Additional documentation in support of the Level II ERA is provided in Appendix A.

Level II (Screening) Ecological Risk Assessment DRAFT Fred Wahl Boatyard July 2010 1

File: X:\0326.01\Projects\03\Fig1_Site Location.mxd

Site

Source: Maul Foster Alongi (2009) Site Address: Tax Lots 400, 500, 601, and 1500, Figure 1-1 Toledo, Lincoln County, Oregon Source: US Geological Survey (1990) 7.5-minute Site Location topographic quadrangle: Toledo South Section 18, Township 11 S, Range 10 W Port of Toledo Toledo, Oregon

0 1,000 2,000 This product is for informational purposes and may not have been prepared for, or be suitable for legal, engineering, or surveying purposes. Users of this information should review or consult the primary data and information sources to ascertain the usability of the information. Feet Project: 0326.01.03 ProducedR. By:Maronn ApprovedM. By: Gibson PrintDate: 08-03-2009 File: X:\0326.01\Projects\03\Fig3_Site Features.mxd X:\0326.01\Projects\03\Fig3_Site File:

Former Paint Mixing/Thinner North of Boatyard Recycle Containment

Service Pier

Former Dry Storage for Vessels

Dry Dock

Closed Loop Washdown/Recycle System Main Building

Travel Lift

Former Paint Storage Containment Floating Moorage

Boatyard Former Parts Sandblasting Containment Yaquina River

Service Pier

Covered Workspace

Floating Moorage

Wetland

Source: adapted from Maul Foster Alongi (2009) Figure 1-2 Aerial photograph (July 2005) obtained from Mapper/USDA; approximate tax lot boundaries Survey Area digitized from figures acquired from Lincoln County Legend Web site (2008) and information provided by Client Port of Toledo Approximate Survey Boundary Note: All locations are approximate. Toledo, Oregon Approximate Habitat Boundary 050100

This product is for informational purposes and may not have been prepared for, or be suitable for legal, engineering, or surveying purposes. Users of this information should review or consult the primary data and information sources to ascertain the usability of the information. Feet Project: 0326.01.03Project: Coffey Maronn/W. R. By: Produced Gibson M. By: Approved 08-29-2009 Date: Print

1.1 SITE HISTORY The Fred Wahl Boatyard covers approximately 19.84 ac. The main structure on the site was previously used as an administrative office and maintenance shop. Other structures include two maintenance canopies, five metal container sheds, a small wooden shed, a travel lift, two service piers, a dry dock, and two floating moorage piers. Historical uses of the property include operation of an over-water shake or shingle mill, a crane operation for installing pilings along local rivers and for dock replacement, and a log storage and rail dump for releasing timber into the Yaquina River. Fred Wahl purchased the property in 1999 and operated a boatyard at the site until 2008, when it was closed down. Currently, the Port of Toledo is interested in purchasing the property from Mr. Wahl and re-opening the boatyard.

1.2 REGULATORY STATUS The Site was the subject of a Phase II environmental site assessment conducted at the boatyard by Maul Foster & Alongi, Inc., in August 2009 (Maul Foster Alongi 2009). Further investigation of the sediment in the vicinity of the dry dock was conducted in February 2010 (Landau 2010). A Level I ERA was conducted in June 2010 (Windward 2010).

1.3 SUMMARY OF LEVEL I ERA RESULTS This section presents a short summary of the results of the Level I ERA conducted at the Fred Wahl Boatyard on June 17, 2010 (Windward 2010).

1.3.1 Contaminants of Interest The contaminants of interest (COIs) at the boatyard are copper and tributyltin (TBT) in the sediment in the immediate vicinity of the dry dock. Copper was detected at concentrations above the ODEQ Level II screening level for marine sediment, and TBT concentrations exceeded the ODEQ Level II screening level for marine sediment and the ODEQ bioaccumulation screening level for marine fish (ODEQ 2001, 2007; Landau 2010; ODEQ 1998).

1.3.2 Potential Ecological Receptors Three distinct habitat areas were identified at the Site:

 A small, nearshore area north of the boatyard with soft, silty substrate and a significant quantity of woody debris, rocks, and chunks of cement slab

 The boatyard area, which had deeper water and hard-packed substrate

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 A small estuarine wetland area with soft, silty substrate along the southern floating moorage pier Ecological receptors that might utilize these habitat areas include benthic invertebrates, fish, birds, and mammals. The invertebrate community at the Site is an estuarine community of limited species diversity typical of mudflats. Four salmon species (Oncorhynchus kisutch, O. keta, O. tshawytscha, and O. mykiss) and both white and green sturgeon (Acipenser transmontanus and A. medirostris) use the Yaquina River in the vicinity of the Site. Coho salmon and green sturgeon are listed as threatened species by either the US Fish and Wildlife Service (USFWS) (2010) and/or the National Oceanic and Atmospheric Administration (NOAA) National Marine Fisheries Service (2010). Chum salmon, Chinook salmon, and steelhead are listed as either sensitive-vulnerable or sensitive-critical by the Oregon Department of Fish and Wildlife (ODFW) (Wilson 2010). Coho salmon are also state listed as sensitive-vulnerable. Sea-run cutthroat trout (Oncorhynchus clarki) are also present in the area. The Site is also expected to support a variety of fish species typical of estuarine channels, such as Pacific staghorn sculpin (Leptocottus armatus), shiner surfperch (Cymatogaster aggregatta), English sole (Parophrys vetulus), and lamprey species (Lampetra spp.) (Simenstad 1983). The Site provides aquatic habitats for numerous birds, including mallard (Anas platyrhynchos), great blue heron (Ardea herodias), and osprey (Pandion Haliaetus) and aquatic mammals, such as river otters (Lontra canadensis).

1.3.3 Primary Exposure Pathways The COIs are imbedded in the sediment adjacent to the dry dock. The primary exposure pathways include direct contact and ingestion of the sediments by benthic invertebrates and fish, and ingestion by invertebrates and fish of prey items exposed to the sediments.

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2 Site Survey

A field survey of the Yaquina River aquatic habitat bordering the Site was conducted on June 17, 2010, as an element of the Level I and II ERAs. A memorandum summarizing the field survey was submitted to the ODEQ on June 28, 2010 (Windward 2010). The survey encompassed the boatyard and aquatic habitats immediately upstream and downstream of the boatyard comprising the habitat assessment area (Figure 1-2). The aquatic and immediate shoreline environments were evaluated for the presence of fish and benthic invertebrates, and for fish and benthic invertebrate habitat quality. Three habitat areas exhibiting changes in habitat characteristics that may influence fish or invertebrates were identified in the habitat assessment area (Figure 2-1). Nearshore substrate at the Site (within 10 to 20 ft of shore) was predominantly silt mixed with gravel and rocks. A significant amount of woody debris was present in the northern part of the Site. All three habitat areas supported an estuarine invertebrate community of limited species diversity, typical of mudflats. The combination of the low-gradient riverbed (i.e., mudflats), the presence of benthic invertebrates, the slow river flows, and the presence of a wetland indicates that habitat along portions of the Site (not including the boatyard) is suitable for potential juvenile salmonid rearing under the conditions observed. The habitat attributes of the boatyard are marginal to low for juvenile salmonid rearing because of the pier structures and steep shoreline. As previously noted, the vicinity of the Site is also expected to support a variety of fish species typical of estuarine channels, such as Pacific staghorn sculpin (L. armatus), shiner surfperch (C. aggregatta), English sole (P. vetulus), and lamprey species (Lampetra spp.) (Simenstad 1983).

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3 Sampling Results and Screening Results and Analysis This section describes the following elements and analyses, which were completed to support the Level II ERA:

 Summary of Site features and Site-specific ecological receptors and exposure factors

 Calculation of environmental concentrations (ECs) for Site-specific chemical contaminants, and screening of Level II ERA sampling results relative to ODEQ screening-level values (SLVs)

 Refined conceptual site model (CSM) integrating findings from the Level II ERA analysis To provide context for the screening and CSM, the ecology of the Site is described as it relates to potential interactions between receptors and chemical contaminants. The organization and content of this section follow ODEQ guidance for the Level II ERA report (ODEQ 1998, 2001).

3.1 SITE DESCRIPTION The Site has been characterized in a Phase II environmental site assessment (Maul Foster Alongi 2009), as well as more recently during additional sediment sampling (Landau 2010) and the Levels 1 and II survey activities described in Section 2. Key aspects of the shoreline CSM are summarized below:

 The nearshore area north of the boatyard is a mudflat with extensive woody debris and numerous rocks and small chunks of cement slab. The width of the mudflat is approximately 10 to 15 ft at a low tide of -0.9 ft. The slope of the mudflat is generally shallow, approximately 5%. The river flow in this area appears to be relatively slow at high tide, based on the amount of woody debris present, and the area might be part of a back eddy that flows into a slough to the north of the Site. Habitat attributes of this area for fish, including habitat for juvenile salmonid rearing, are marginal to low because of the large amount of wood waste.

 At the boatyard, the nearshore area consists of a narrow strip of muddy substrate with pilings, rocks, gravel, and a few small chunks of cement slab. Part of the shoreline under the service piers is shored up. The width of the muddy shoreline is approximately 5 to 8 ft at a low tide of -0.9 ft. The slope of the narrow shoreline is moderate (estimated at 10 to 15%). The river channel reaches a water depth of approximately 8 to 10 ft along the mooring pier, approximately 30 ft from the

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shore (estimated slope 25 to 35%). The current along the outside of the dry dock and service piers is fast, with an estimated speed of approximately 1.5 m/s during ebb. Habitat attributes for juvenile salmonid rearing in this area are marginal to low because of the pier structures and steep shoreline; however, other fish species are likely to use this area.

 A small, tidally influenced estuarine wetland is located in the southern part of the Site along the shoreline of the Yaquina River. The wetland is approximately 400 ft long with a width that ranges from a few feet to between 20 and 30 ft. Dominant plants are obligate wetland species that typically occur in coastal estuaries on silt and mudflats. The grassy wetland area is adjacent to the exposed intertidal mudflat, suggesting that the intertidal marsh slope is shallow (estimated at 0 to 5%). Woody debris is observable throughout the area. Habitat attributes of this area for fish, including habitat for juvenile salmonid rearing, are high because of the wetland’s presence next to a larger mudflat.

3.2 SITE-SPECIFIC ECOLOGICAL RECEPTORS The target ecological receptors for the aquatic assessment are juvenile salmonids, sturgeon, benthic fish populations, and the benthic invertebrate community. The benthic community includes both infaunal and epibenthic invertebrates; however, the principal invertebrates of interest are the infaunal community living in the substrate of the shoreline area along the Site. The rationale for the selection of ecological receptors is presented in Table 3-1 and discussed in greater detail below.

Table 3-1. Aquatic habitats, supported species, and selected ecological receptors

Habitat Type Extent of Habitat Species Supported Site-Specific Receptor(s) migrating adult salmon; juvenile salmon Nearshore mudflat with small fish (including extending 10-15 ft into river (representing other small extensive woody debris rearing juvenile from shoreline fish); benthic invertebrate and numerous rocks salmonids); benthic community invertebrates Narrow, muddy shoreline with moderate slope muddy shoreline, (10-15%) adjacent to the approximately 5-8 ft. migrating adult salmon; juvenile salmon; sturgeon; river channel. River sturgeon; estuarine fish small benthic fish channel with steeper species; benthic populations; benthic slope (25-35%) and a invertebrates invertebrate community current speed of river channel, 100-150 ft approximately 1.5 m/s during ebb. Small, tidally influenced wetland approximately 400 ft migrating adult salmon; juvenile salmon estuarine wetland long with a width of 20-30 ft small fish (including (representing other small

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Habitat Type Extent of Habitat Species Supported Site-Specific Receptor(s) adjacent to intertidal rearing juvenile fish); benthic invertebrate mudflat mudflat extending 15-20 ft salmonids); benthic community from the wetland invertebrates

3.2.1 Salmonids As a conservative supposition, this assessment assumes that juvenile salmon listed as either threatened under the Endangered Species Act (ESA) of 1973 (coho), or as sensitive-critical/vulnerable by ODFW (chum, Chinook, and steelhead), could be present in the Level II ERA study area (although juvenile salmon have not been observed there). No habitat suitable for salmonid spawning was observed at the Site, so eggs and early life stages would not occur. Out-migrating juvenile salmon are likely to rear in the small wetland and the mudflat at the Site. The likelihood that the boatyard and the northern nearshore area support juvenile salmon habitats is low, because of the large amount of wood waste and the presence of overwater structures. Wood waste limits soft substrates for burrowing organisms, can create anoxic conditions during certain times of year because of high biological oxygen demand, and can be associated with toxic leachates and hydrogen sulfide (Toews and Brownlee 1981). Juvenile salmon typically will avoid the shadows created by overwater structures, likely because of impaired visual sensitivity (Nightingale and Simenstad 2001). Adult salmon may utilize the aquatic habitats at the Site during migration. Chinook, coho, and chum spawn in freshwater during the fall (WDFW 2010). Chinook migrate furthest up the watershed, whereas chum generally spawn closer to salt water. The Yaquina River has both summer and winter runs of steelhead. Steelhead spawn in freshwater during the spring and closer to the ocean during the winter. The juveniles of all these species, with the exception of chum, stay in freshwater for at least a year. Chum fry do not rear in freshwater for more than a few days; shortly after they emerge, chum fry move downstream to the estuary and rear there for several months before heading out into the open ocean. To varying degrees, juvenile salmon depend on estuarine habitats for rearing during their out-migration from freshwater to salt water. Juvenile Chinook salmon are generally regarded as the most estuarine dependant, with residence times of several months (Healy 1991). Chum are also highly estuarine dependant, residing there for anywhere from several days to months (Salo 1991). Coho and steelhead do not have extensive estuarine residence times (Sandercock 1991). Juvenile salmon shift their estuarine habitat use with tidal cycles and over time with growth (Healy 1991), so they are likely to forage over an extended area during their estuarine residences.

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3.2.2 Sturgeon This assessment assumes that green sturgeon listed as a threatened species by either the USFWS (2010) and/or the NOAA National Marine Fisheries Service (2010) could be present in the Level II ERA study area. Both green and white sturgeon might utilize the aquatic habitats at the Site. Sturgeon rely on large, complex river systems for many of their life stages and can feed opportunistically on prey ranging from benthic invertebrates to large fish (Beamesderfer and Farr 1997). Both types of sturgeon have very long life spans (e.g., 70-100 years) (Dees 1961; Nakamoto et al. 1995). Sturgeon life history can be divided into three general phases: freshwater juveniles (<3 years), coastal migrants, and adults (≥ 13 years for females, ≥ 9 years for males) (Nakamoto et al. 1995; Wydoski and Whitney 2003). Some studies suggest that sturgeon can show strong site fidelity (Veinott et al. 1999), while other studies indicate that individual sturgeon have wide site ranges (DeVore and Grimes 1993). The home ranges of sturgeon are studied through the use of passive integrated transponder (PIT) tags or spaghetti wires, which are attached to sturgeon that are caught and released to track their movements. One such study recorded a juvenile white sturgeon migrating approximately 72 mi between the tagging location and where the sturgeon was collected (WDFW 2007). The movements of this fish indicate a wide home range for sturgeon, even during their pre-breeding life stage.

3.2.3 Invertivorous fish Invertivorous (benthic) fish populations, such as sculpins and flatfishes (e.g., English sole), might utilize the benthic habitats at the Site (Simenstad 1983). Other species, such as shiner surfperch or pile perch, might use the pilings as an artificial reef. Pacific staghorn sculpin and shiner surfperch are among the most common fish in Oregon estuaries, including the Yaquina River estuary (Bayer 1985). Pacific staghorn sculpin are benthic fish that prefer sandy areas (Eschmeyer et al. 1983). Sculpins live approximately 4 to 5 years and reach maturity within 2 to 4 years. They are benthic feeders as adults and consume crustaceans, aquatic insect larvae, fish, and mollusks. The foraging range of Pacific staghorn sculpin is unknown. However, based on low variability in tissue chemical concentrations throughout the Lower Duwamish Waterway Superfund site (where there is high variability in sediment chemical concentrations), foraging ranges are probably more than 1 mile (Windward 2007). Shiner surfperch are opportunistic carnivores found in loose schools near piers and pilings. They feed on zooplankton, small crustaceans, algae, and detritus, as well as polychaetes, mollusks, and benthic organisms (Gordon 1965; Bane and Robinson 1970). Pile perch use similar habitats and feed on hard-shelled mollusks, as well as

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crustaceans, such as barnacles, and crabs (Eschmeyer et al. 1983). The foraging range of shiner surfperch is also unknown. However, based on low variability in tissue contaminant concentrations throughout the Lower Duwamish Waterway Superfund site, foraging ranges are probably more than 1 mile (Windward 2007). English sole are typically found on soft sand or mud bottoms (Eschmeyer et al. 1983). Juvenile English sole settle in estuarine areas and migrate to deeper water habitats as adults. Juveniles consume annelid worms, copepods, amphipods, and mollusks. Adult English sole ingest clams, marine worms, small crabs, and small shrimps (Fresh et al. 1979). English sole are thought to have limited foraging areas. Based on a review of the available data, Stern et al. (2003) suggest a home range of approximately 2 km2.

3.2.4 Benthic invertebrate community The June 2010 Site survey observed clams (Macoma sp.), amphipods, polychate, barnacles, and mysid shrimps. Based on these observations, the invertebrate community at the Site can be described as an estuarine community typical of mudflats. In rivers, benthic invertebrates are generally limited to a narrow band of habitat near the substrate surface. In fine-grained sediments, invertebrates are generally limited to the top 10 cm because they require surface water for ventilation (SEA 2002; Boudreau 1998). The decomposed organic material that settles to the bottom of the river consumes dissolved oxygen, creating an anoxic environment, and releases products that result in the formation of hydrogen sulfide and ammonia. These conditions create a toxic environment for infaunal organisms, thus preventing invertebrates from living at greater depths in the fine-grained sediments (Forbes et al. 1998; Hershey and Lamberti 1998). Invertebrates that live in the sediment (infaunal invertebrates) are exposed to contaminants via dermal contact and the porewater that they respire. Porewater generally refers to interstitial water in fine-grained sediments. Infaunal invertebrates meet their need for oxygen through modifications to the sediment environment, or through behaviors that increase the interchange with overlying water, thereby decreasing their exposure to porewater contaminants. Mechanisms and behaviors that alter an infaunal organism’s exposure to porewater vary by species. Burrowing and tube-dwelling organisms may actively pump overlying water into their burrows or tubes through regular beating of various structures, such as pleopods (e.g., crayfish) or cilia, or through body undulations or peristaltic contractions (e.g., some soft-bodied worms) (Riisgard and Larsen 2005). Some infauna construct U-shaped burrows or tubes with one opening slightly higher in elevation than the other. This difference in height creates a passive flow-through system that minimizes the metabolic energy required to flush their tubes or burrows (Vogel 1994). The entrainment of overlying water into

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tubes and burrows oxygenates not only the tube or burrow, but also the sediment surrounding the tube or burrow (Satoh et al. 2007). All invertebrates may also be exposed to chemicals through feeding. Filter-feeding organisms are dependent upon the food brought to them by flowing water. Organisms that feed in this manner extend specialized appendages or structures into the water column to gather food. As an example, a filter-feeding bivalve will extend its siphon above the sediment surface and pump overlying water via gill ciliary action across the gills, through the mantle cavity, and out the siphon. This action limits exposure to porewater while supporting both feeding and respiration. Macoma clams, for example, are estimated to ventilate approximately 10% porewater, even when their siphons are retracted inside their burrows (Winsor et al. 1990). Infauna that feed on organic material below the sediment surface tend to increase the porosity of the sediments by their activities, which increases the exchange of porewater with overlying water (Krantzberg 1985; Winsor et al. 1990).

3.3 CONTAMINANTS OF POTENTIAL ECOLOGICAL CONCERN Copper and TBT are the sediment COIs identified in the Level I ERA and are the COIs for the Level II ERA. This section describes the comparison of the sediment results to the screening values in order to identify contaminants of potential ecological concern (CPECs) and support the Level II ERA evaluations. Sediment samples were collected primarily to support the characterization of the nature and extent of contamination (Landau 2010; Maul Foster Alongi 2009). Sampling results for copper were compared to ODEQ Level II ERA marine sediment SLVs to identify locations with elevated concentrations (Table 3-2). The copper concentrations were screened by dividing the maximum concentration in any sample by the sediment SLV to calculate an exceedance factor (EF) according to ODEQ guidance (ODEQ 1998, 2001).

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Table 3-2. Sediment screening values

ODEQ SLVa Other SLV Marine Fish Analyte Unit Marine Bioaccumulationb Meadorc SQuiRTd Copper mg/kg dw 19e nv nv 180f

TBT µg/kg dw nvh 0.37 120 > 1,391g

a ODEQ Eco SLV – Guidance for Ecological Risk Assessment: Levels I, II, III, IV, Table 2, Level II Screening Level Values for Freshwater and Marine Sediment (ODEQ 1998, 2001) b DEQ Guidance for assessing bioaccumulative chemicals of concern in sediment, Table A-1a, Sediment Bioaccumulation Screening Level Values (SLVs) (DEQ 2007) c Meador et al. (2002) d SQuiRTs (NOAA 2008) e Based on TEL, marine sediment f Based on PEL, marine sediment g Based on AET, marine sediment as listed in SquiRTs (NOAA 2008). h Incorrect value presented in the ODEQ guidance (ODEQ 1998, 2001) AET – apparent effects threshold PEL – probably effects level DEQ – Department of Environmental Equality SLV – screening-level value dw – dry weight SQuiRT – Screening Quick Reference Table nv – no value available for this compound TBT – tributyltin ODEQ – Oregon Department of Environmental Quality TEL – threshold effects level The sediment screening value for copper is based on a threshold effects level (TEL) for marine sediments (ODEQ 1998, 2001). The TELs are intended to estimate the concentrations of chemicals below which adverse biological effects only rarely occur (Smith et al. 1996). The TELs are derived by calculating the geometric mean of the 15th percentile of the effect dataset and the 50th percentile of the no-effect dataset. For perspective, copper sediment data are compared to the SQuiRT probable effects level (PEL) for marine sediments. The PELs are intended to estimate the concentration of a chemical above which adverse biological effects frequently occur (Smith et al. 1996). The PELs are derived by calculating the geometric mean of the 50th percentile of the effect dataset and the 85th percentile of the no-effect dataset.

The ODEQ Level II ERA marine sediment TBT SLV as presented in the guidance document is in error. According to ODEQ, (1998, 2001) the sediment screening value for TBT is based on an apparent effects threshold (AET) for marine sediments as derived in the SQuiRT tables by NOAA (2008). AET is the sediment contaminant concentration above which statistically significant (p ≤ 0.05) adverse biological effects (relative to appropriate reference conditions) would always be expected (Gries and Waldow 1996). In the ODEQ guidance, an SLV of 3 µg/kg dry weight (dw) is listed, but no

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corresponding marine sediment value is listed in the SQuiRT table for TBT or tributyltinoxide (NOAA 2008). Using the marine sediment value of > 3,400 µg/kg dw listed in the SQuiRT table for tin and a conversion factor of 2.444 (EPA 2003), an SLV of > 1,391 µg/kg dw can be derived for TBT. The ODEQ bioaccumulation screening value for marine fish is based on a conversion of the marine Ambient Water Quality Criterion for TBT into a sediment value using a bioconcentration factor and a biota-sediment accumulation factor. The ODEQ marine fish bioaccumulation SLV represents a chemical concentration in sediment at and below which the chemical would not be expected to accumulate in the tissues of fish or other aquatic organisms above levels acceptable to those organisms (ODEQ 2007), even if conditions were favorable for bioaccumulation. Another pertinent sediment SLV for TBT is a screening value derived by the National Marine Fisheries Service (NMFS) (Meador et al. 2002). This TBT SLV is intended to be protective of salmonid prey items and is based on a tissue residue concentration that has been correlated with reduced growth in at least seven salmonid prey items and an organic carbon-normalized bioaccumulation factor that represents relatively sensitive species. The resulting sediment value of 6,000 ng/g organic carbon is converted to 120 µg/kg dw based on a total organic carbon (TOC) value of 2%. Copper was identified as a CPEC because sediment concentrations at five locations exceeded the marine SLV (Table 3-3). When compared to the PEL for marine sediments, none of the concentrations or the upper confidence limit (UCL) exceeded the SLV (the UCL calculations are provided in Appendix A). As stated above, the narrative intent of TEL is to estimate the concentration below which adverse biological effects only rarely occur, whereas the narrative intent of PEL is to estimate the concentration of a chemical above which adverse biological effects frequently occur.

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Table 3-3. Sediment sampling results compared to SLVs

EF (unitless) EC Copper Tributyltin Copper Tributyltin (mg/kg (µg/kg ODEQ Marine Fish Sampling Location dw) dw) Marinea SQuiRTb Bioaccumulationb SQuiRTc Meadord SED-1 82.1 64 4.3 0.5 173.0 < 0.046 0.5 SED-2 15.2 2.1U 0.8 0.1 5.7 < 0.002 0.0 1A 36 110 1.9 0.2 297.3 < 0.079 0.9 2A 18.7J 3.4U 1.0 0.1 9.2 < 0.002 0.0 3A 8.4 3.4U 0.4 0.0 9.2 < 0.002 0.0 4A 5.3 3.7U 0.3 0.0 10.0 < 0.003 0.0 5A 80.4 3.7U 4.2 0.4 10.0 < 0.003 0.0 6A 92.9 3.4U 4.9 0.5 9.2 < 0.002 0.0 7B 22.5 11 1.2 0.1 29.7 < 0.008 0.1 9B 5.9 3.6U 0.3 0.0 9.7 < 0.003 0.0 UCL 80.7 110 4.2 0.4 297.3 < 0.079 0.9 Average 39.1 21.7 2.1 0.2 58.6 < 0.016 0.2 a ODEQ Eco SLV – Guidance for Ecological Risk Assessment: Levels I, II, III, IV, Table 2, Level II Screening Level Values for Freshwater and Marine Sediment (ODEQ 1998, 2001) b DEQ Guidance for assessing bioaccumulative chemicals of concern in sediment, Table A-1a, Sediment Bioaccumulation Screening Level Values (SLVs) (DEQ 2007) c SQuiRT (NOAA 2008). d Meador et al. (2002). dw – dry weight ODEQ – Oregon Department of SQuiRT – Screening Quick EC – environmental concentration Environmental Quality Reference Table (sample result) SLV – screening level value U – undetected at detection limit EF – exceedance factor (EF = shown EC/SLV) UCL - upper confidence limit J – estimated concentration

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TBT was identified as a CPEC because detected sediment concentrations at three of the nine sampling locations exceeded the ODEQ bioaccumulation screening value for marine fish (ODEQ 1998, 2001). The bioaccumulation SLV represents the chemical concentration in sediment at and below which the possibility of the chemical accumulating in tissues of fish or other aquatic organisms to a level that would pose a risk can be confidently screened out, even without conducting a full risk assessment. The Meador et al. (2002) TBT SLV is protective of salmonid prey items and is derived using a TOC value of 2%. None of the sediment TBT concentrations measured at the Site exceed this value. TOC has a major effect on the bioavailability of TBT and therefore the toxicity and bioaccumulation of TBT. Another study by Meador et al. shows that as the TOC level increases, the tissue residues and toxicity decrease (Meador et al. 1997). TOC values as high as 17.2% have been measured at the Site (Maul Foster Alongi 2009); hence the Meador et al. value is a conservative sediment value to screen TBT risk at the Site. The other TBT SLV that is available—NOAA’s SQuiRT value for tin, converted to TBT using the US Environmental Protection Agency’s (EPA’s) conversion factor of 2.444 (EPA 2003)—is an order of magnitude higher than the Meador et al. SLV. None of the sediment samples at the Site exceeded the Meador et al. SLV or the converted SQuiRT value.

3.4 PRELIMINARY CONCEPTUAL SITE MODEL This section discusses the preliminary CSM integrating information on ecologically important receptors, assessment endpoints, CPECs, exposure pathways, and potential effects.

3.4.1 Known ecological effects This section discusses chemical-specific factors (from literature) that affect the toxicity and behavior of copper and TBT with respect to fish and benthic invertebrates.

Copper Copper is the active agent in many antifouling bottom paints applied to boats. The use of copper-based paints increased following the 1982 prohibition of TBT-based paints (Hall et al. 1988; as cited in Eisler 1997). Sediment interstitial porewater copper concentrations correlate positively with dissolved copper concentrations in the overlying water column (Ankley et al. 1993; as cited in Eisler 1997). In marine sediments, the bioavailability of copper increases with increasing acid volatile sulfide content (Casas and Crecelius 1994; as cited in Eisler 1997).

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In the aqueous environment, copper toxicity and bioavailability are influenced by a variety of water quality parameters, rather than simply by the total concentration of copper in the water column. The biotic ligand model describes toxicity as a function of free ions of copper in the aqueous environment forming complexes with a variety organic and inorganic ligands (e.g., dissolved organic carbon), including the site of toxic action within the organism, the biotic ligand (e.g., gill), and the competition for binding sites on the biotic ligand with a variety of cations (e.g., Ca2+, Na+, and H+). Toxicity occurs when the copper concentration on the biotic ligand exceeds a threshold, thus disrupting a critical biological function. For fish, the biotic ligand appears to correspond to sites on the surface membrane of the gill responsible for regulating sodium ion uptake. Osmoregulation is disrupted by inhibition of the enzyme Na, K ATPase in the gill tissue and thereby the uptake of plasma electrolytes (Heath 1995). In fish, growth and survival are generally the endpoints at which organismal-level toxicity is observed (Erickson et al. 2003; Mount et al. 1994; Handy 1992). Copper is acutely toxic to invertebrates with bivalve embryos among the most sensitive species (EPA 1985). In gastropods, exposure to high sublethal copper concentrations adversely affects surface epithelia, disrupting osmoregulation and producing water accumulation in tissues (Cheng 1979 ; as cited in Eisler 1997). It should be noted that gastropods were not observed at the Site during the survey.

TBT TBT was formerly used in antifouling bottom paints for boats, so barnacles and other fouling organisms are often exposed to its acute toxicity(EPA 2003). TBT has a log octanol-water partition coefficient (KOW) of 4.4 and partitions between water, sediment, and organisms in a manner somewhat consistent with equilibrium partitioning (Meador 2000). Therefore, organic carbon components of sediment and water and lipid content of the organism strongly affect bioavailability (Meador 2000). However, Meador (2000) reports that dry weight bioconcentration factors (BCFs) range from approximately 1,000 for Daphnia magna to 900,000 for a freshwater bivalves. These values vary considerably from the predicted BCF of 5,800, which was based on equilibrium partitioning, indicating that bioaccumulation of TBT strongly depends on species-specific toxicokinetics (Meador 2000). Thus, generic BCFs cannot be used to reliably predict tissue burdens in aquatic organisms. Acute toxicity of TBT appears to be caused by uncoupling of oxidative phosphorilation (Fent 1996), whereas an important mode of action for sublethal effects is endocrine disruption (Meador 2000). Snails have been identified as the most sensitive species based on the imposex endpoint (superimposition of male sexual characteristics on female snails) (Gibbs et al. 1988). Larval bivalve mollusks and juvenile crustaceans appear to be much more sensitive than adults during acute exposures (EPA 2003). Disappearance of clam populations, including Macoma baltica and Scrobicularia plana, Level II (Screening) Ecological Risk Assessment DRAFT Fred Wahl Boatyard July 2010 17

have been reported at sediment concentrations exceeding 700 and 800 µg/kg dw (Meador et al. 2002). TBT has been identified as a bioaccumulative compound potentially causing adverse effects on organisms higher up in the food chain (EPA 2003). In fish, exposure to TBT has been shown to cause adverse effects on a variety of reproductive endpoints, including fertilization success and survival of early life stages from eggs through juveniles (Shimasaki et al. 2006; Nakayama et al. 2005; Nirmala et al. 1999), and on sexual differentiation and growth (Shimasaki et al. 2003).

3.4.2 Candidate assessment endpoints The candidate assessment endpoint for juvenile salmonids is maintenance of individual juvenile salmon survival and growth. An individual endpoint, rather than one for population-level assessment, is appropriate because coho salmon present in the Yaquina River is a threatened species under the ESA. Reproduction of salmonids is an inappropriate assessment endpoint because salmonids do not spawn in the vicinity of the Site, and returning adult salmonids would only briefly be present during migration; therefore, reproductive life stages (adults and eggs) are not expected to be exposed to Site contaminants. The assessment endpoint for green sturgeon is maintenance of individual survival and growth. An individual endpoint, rather than one for population-level assessment, is appropriate because green sturgeon is a threatened species under the ESA. Reproduction of green sturgeon is an inappropriate assessment endpoint because sturgeon do not spawn in the vicinity of the Site. The assessment endpoint for white sturgeon and invertivorous fish is maintenance of survival and growth of the population. For small, home-range invertivorous fish, such as sculpins, an additional assessment endpoint is maintenance of reproduction of the population. An appropriate assessment endpoint for benthic invertebrates is maintenance of survival, growth and/or reproduction of the benthic community, and maintenance of community structure.

3.4.3 Relevant and complete exposure pathways This section discusses the potential for receptors in the Yaquina River to have significant exposure to CPECs. The interpretations presented are based on the screening results presented in Section 3.3, and the CSM developed for the shoreline area and Level II screening analysis. The exposure pathway determinations for the shoreline area were made following the criteria presented in technical management decision point (TMDP) 3 Ecological Risk Probable?, of the ODEQ Level II screening guidance (ODEQ 1998, 2001).

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For CPECs to pose a potential risk to receptors, the exposure pathway must be complete. An exposure pathway is considered complete if a chemical can migrate from its source to an ecological receptor, and the receptor is exposed via one or more exposure routes at concentrations above risk-based SLVs. Complete exposure pathways can be of varying importance, so key pathways that reflect maximum exposures of ecological receptors are identified as having more importance than pathways likely to provide a very low fraction of the receptors’ total exposure to CPECs. Pathways for the exposure of receptors to CPECs were designated in one of three ways: complete and significant, complete and insignificant, or incomplete. Each of these three designations is defined below, including whether the pathway warrants further evaluation in a higher-tiered risk assessment. This section also presents a brief rationale for each exposure pathway designation. The CSM is presented in Figure 3-1.

 Complete and significant. This pathway is identified when there is a direct link between the receptor and the chemical via the pathway, and the specific pathway is considered to be potentially important. Pathways classified as complete and significant are recommended for further risk assessment.

 Complete and insignificant. This pathway is identified when there is a direct link between the receptor and the chemical via the pathway, but the significance of this pathway in terms of exposure is considered to be very low. Pathways classified as complete and insignificant are not recommended for further risk assessment.

 Incomplete. This pathway is identified when there is no direct pathway between the receptor and the chemical at concentrations above SLVs. Pathways classified as incomplete are not recommended for further evaluation.

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Figure 3-1. Conceptual site model

Fish Juvenile salmon Direct sediment contact, sediment ingestion, prey ingestion, and indirect contact via surface water are complete and insignificant exposure pathways for juvenile salmon. Juvenile salmon reside in the water column, where they forage mainly on drifting insects and to a lesser extent on benthic invertebrates (Healy 1991). They do not have appreciable amounts of sediment in their stomachs (Cordell 2001) so, juvenile salmon are not likely to be significantly exposed to contaminants through sediment ingestion. The prey ingestion pathway is also complete and insignificant. For juvenile salmon, the habitat attributes of the area with elevated copper and TBT concentrations are marginal to low because of the pier structures and steep shoreline. Furthermore, juvenile salmon grow rapidly during estuarine residence and shift their habitat use with growth (Healy 1991), so they do not reside for extended periods in the small area of contaminated sediment along the Site. Thus, because of their diet (obtaining a majority of their food from the water column), their habitat use (pelagic), the limited area of affected sediment at the boatyard, and the low habitat value of this area, juvenile salmon are unlikely to

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have significant exposure to the Site CPECs in sediment, either directly or through the food web. Direct sediment contact, sediment ingestion, prey ingestion, and indirect contact via surface water are complete and insignificant exposure pathways for adult salmonids. Adult salmon may briefly utilize the habitats at the Site as a migratory corridor. However, the time that adult salmon spend under the dry dock and piers is likely very small relative to the area over which adult salmon migrate. Sturgeon Direct sediment contact, sediment ingestion, prey ingestion, and indirect contact via surface water are complete and insignificant exposure pathways for both green and white sturgeon. The surface area of sediments along the Site with elevated TBT and copper concentrations is very small relative to the area over which sturgeon forage in the river environment. Sturgeon that may utilize the habitats at the Site would likely be young adults or adults (> 3years). At these stages of their lives, sturgeon primarily feed opportunistically on prey ranging from benthic invertebrates to large fish (Beamesderfer and Farr 1997). The proportion of exposure via all pathways in this area of the Site is certain to be very small compared to the overall foraging area. Invertivorous fish Direct sediment contact, sediment ingestion, prey ingestion, and indirect contact via surface water are complete and insignificant pathways for invertivorous fish. Invertivorous fish, such as Pacific staghorn sculpin and shiner surfperch, likely consume benthic invertebrates at the Site and respire overlying surface water. During their residence at the site, invertivorous fish could be directly exposed to contaminants in surface sediment, or indirectly exposed through prey or surface water. However, invertivorous fish foraging ranges are probably more than 1 linear mile of estuary, so the Site likely constitutes a small fraction of their overall exposure. Furthermore, although individual fish may be exposed, the small size of the site relative to the area over which breeding populations likely extend suggests that any exposure at the Site would not affect fish populations.

Benthic invertebrate community Direct sediment contact, sediment ingestion, and prey ingestion are complete and significant pathways for the benthic invertebrate community. Based on screening level assumptions, it is possible that benthic invertebrates living in the area adjacent to the dry dock might be adversely affected by the localized sediment copper and TBT concentrations. However, because the assessment endpoint is maintenance of survival, growth and/or reproduction of the benthic community, and maintenance of community structure, adverse effects to a limited number of invertebrates would not constitute adverse effects to the benthic community at the Site. Indirect contact via Level II (Screening) Ecological Risk Assessment DRAFT Fred Wahl Boatyard July 2010 21

surface water is a complete and insignificant exposure pathway for the benthic invertebrate community.

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4 Conclusions and Recommendations

This section presents the conclusions and recommendations of the Level II ERA based on analysis of the ecology of the shoreline environment, chemical data collected, and potential for exposure of ecological receptors to COIs under the Site-specific conditions. The primary findings of this Level II ERA are as follows:

 The concentrations of copper at five sediment locations adjacent to the dry dock exceed the ODEQ SLV of 19 mg/kg dw, which is based on a TEL with the narrative intent of estimating the concentration below which adverse biological effects only rarely occur. None of the concentrations or the UCL exceed the PEL with the narrative intent of estimating the concentration of a chemical above which adverse biological effects frequently occur. Hence, the small extent of contamination combined with the itinerant habitat use and relatively large foraging ranges of the fish suggest that copper poses no risks to ecological receptors at the Site.

 TBT was identified as a CPEC because detected sediment concentrations at three of the nine sampling locations exceeded the ODEQ TBT bioaccumulation screening value for marine fish (ODEQ 2007). The sediment TBT concentrations did not exceed the marine AET value (NOAA 2008) or the sediment value by Meador et al (2002). Both of these sediment values are conservative values protective of marine organisms and salmonid prey items. The small extent of contamination combined with the itinerant habitat use and relatively large foraging ranges of the fish suggest that bioaccumulation from sediment would not result in any adverse effects on individual juvenile salmon or populations of other fish receptors. The findings of this Level II ERA show that there is no unacceptable risk to the aquatic ecological receptors present in the aquatic habitats at the Fred Wahl boatyard. Based on these findings, and in accordance with TMDP 3 of the ODEQ Level II screening guidance (ODEQ 1998, 2001), no further ecological investigation is necessary to reach a no-risk conclusion for the Site.

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5 References

Ankley GT, Mattson VR, Leonard EN, West CW, Bennett JL. 1993. Predicting the acute toxicity of copper in freshwater sediments: evaluation of the role of acid-volatile sulfide. Environ Toxicol Chem 12:315-320. Bane G, Robinson M. 1970. Studies on the shiner perch, Cymatogaster aggregata Gibbons, in upper Newport Bay, California. Wasmann J Biol 28(2):259-268. Bayer RD. 1985. Shiner perch and Pacific staghorn sculpins in Yaquina Estuary, Oregon. Northw Sci 59(3):230-240. Beamesderfer RCP, Farr RA. 1997. Alternatives for the protection and restoration of sturgeons and their habitat. Environ Biol Fish 48:407-417. Boudreau BP. 1998. Mean mixed depth of sediments: the wherefore and the why. Limnol Oceanogr 43(3):524-526. Casas AM, Crecelius EA. 1994. Relationship between acid volatile sulfide and the toxicity of zinc, lead and copper in marine sediments. Environ Toxicol Chem 13(3):529-536. Cheng TC. 1979 Use of copper as a molluscicide. In: Nriagu JO, ed, Copper in the environment. Part 2: health effects. John Wiley, New York, NY, pp 401-432. Cordell J. 2001. Personal communication (e-mail to Matt Luxon, Windward Environmental, Seattle, WA, regarding observations of juvenile chinook and other wildlife in the Lower Duwamish River). Researcher, Department of Fisheries, University of Washington, Seattle, WA. July 9. Dees LT. 1961. Sturgeons. Fishery leaflet #526. US Fish and Wildlife Service, Washington, DC. DEQ. 2007. Guidance for assessing bioaccumulative chemicals of concern in sediment. 07-LQ-023A. Environmental Cleanup Program, Oregon Department of Environmental Quality, Portland, OR. DeVore JD, Grimes JT. 1993. Migration and distribution of white sturgeon in the Columbia River downstream from Bonneville Dam and in adjacent marine areas. In: Beamesderfer RC, Nigro AA, eds, Status and habitat requirements of the white sturgeon populations in the Columbia River downstream from McNary Dam. Final report of research. Vol I. Prepared for Bonneville Power Administration. Oregon Department of Fish and Wildlife, Portland, OR, pp 83- 100.

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Eisler R. 1997. Copper hazards to fish, wildlife and invertebrates: a synoptic review. Patuxent Wildlife Research Center, US Fish and Wildlife Service, Laurel, MD. EPA. 1985. Ambient water quality criteria for copper-1984. EPA 440/5-84-031. Office of Water, US Environmental Protection Agency, Washington, DC. EPA. 2003. Ambient aquatic life water quality criteria for tributyltin (TBT) - final. EPA 822-R-03-031. US Environmental Protection Agency, Washington, DC. Erickson RJ, Highland TL, Hocket JR, Leonard EN, Mattson VR, Mount DR. 2003. Effects of dietary copper, zinc, lead, cadmium, and arsenic on growth and survival of juvenile fish using live food organisms. Platform presentation at SETAC 24th annual meeting, Austin TX, 9-13 November 2003. Manuscript in prep. Eschmeyer WN, Herald ES, Hammann H. 1983. Pacific coast fishes. Peterson Field Guide Series. Houghton Mifflin, Boston, MA. Fent K. 1996. Ecotoxicology of organotin compounds. Crit Rev Toxicol 26(1):1-117. Forbes TL, Forbes VE, Giessing A, Hansen R, Kure LK. 1998. Relative role of pore water versus ingested sediment in bioavailability of organic contaminants in marine sediments. Environ Toxicol Chem 17(12):2453-2462. Fresh KL, Rabin D, Simenstad CA, Salo EO, Garrison K, Matheson L. 1979. Fish ecology studies in the Nisqually Reach area of southern Puget Sound, Washington. FRI- UW-7904. Prepared for Weyerhauser Company. Fisheries Research Institute, University of Washington, Seattle, WA. Gibbs PE, Pascoe PL, Burt GR. 1988. Sex change in the female dog-whelk, Nucella lapidis, induced by tributyltin from antifouling paints. J Mar Biol Ass UK 68:715-731. Gordon CD. 1965. Aspects of the life-history of Cymatogaster aggregata Gibbons. MS thesis. University of British Columbia, Victoria, BC. Gries TH, Waldow KH. 1996. Progress re-evaluating Puget Sound apparent effects thresholds (AETs). Volume I: 1994 amphipod and echinoderm larval AETs. Draft report. Prepared for Puget Sound Dredged Disposal Analysis (PSDDA) agencies. Washington Department of Ecology, Olympia, WA. Hall LW, Jr,, Bushong SJ, Hall WS, E JW. 1988. Acute and chronic effects of tributyltin on a Chesapeake Bay copepod. Environ Toxicol Chem 7:41-46. Handy RD. 1992. The assessment of episodic metal pollution. II. The effects of cadmium and copper enriched diets on tissue contaminant analysis in rainbow trout (Oncorhynchus mykiss). Arch Environ Contam Toxicol 22:82-87.

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Healy MC. 1991. Life history of chinook salmon (Oncorhynchus tshawytscha). In: Groot C, Margolis L, eds, Pacific salmon life histories. UBC Press, Vancouver, BC, pp 311- 394. Heath AG. 1995. Water pollution and fish physiology. Second ed. CRC Press Inc., Boca Raton, FL. Hershey AE, Lamberti GA. 1998. Stream macroinvertebrate communities. In: Naiman RJ, Bilby RE, eds, River ecology and management: lessons from the Pacific coastal ecoregion. Springer, New York, NY, pp 169-1999. Krantzberg G. 1985. The influence of bioturbation on physical, chemical and biological parameters in aquatic environments. A review. Environ Pollut (Series A) 39(2):99-122. Landau. 2010. Fred Wahl Shipyard phase II sediment investigation report, Port of Toledo, Toledo, Oregon. Landau Associates, Portland, OR. Maul Foster Alongi. 2009. Phase II environmental site assessment, Fred Wahl Marine Construction site, 621 and 1000 Altree Land, Toledo, Oregon. . Maul Foster Alongi (MFA), Portland, OR. Meador JP. 2000. Predicting the fate and effects of tributyltin in marine systems. Rev Environ Contam Toxicol 166:1-48. Meador JP, Krone CA, Dyer DW, Varanasi U. 1997. Toxicity of sediment-associated tributyltin to infaunal invertebrates: species comparison and the role of organic carbon. Mar Environ Res (43):219-241. Meador JP, Collier TK, Stein JE. 2002. Determination of a tissue and sediment threshold for tributyltin to protect prey species of juvenile salmonids listed under the US Endangered Species Act. Aquat Conserv Mar Freshw Ecosys 12:539-551. Mount DR, Barth AK, Garrison TD, Barten KA, Hockett JR. 1994. Dietary and waterborne exposure of rainbow trout (Oncorhynchus mykiss) to copper, cadmium, lead and zinc using a live diet. Environ Toxicol Chem 13(12):2031-41. Nakamoto RJ, Kisanuki TT, Goldsmith GH. 1995. Age and growth of Klamath River green sturgeon (Acipenser medirostris). Klamath River Fishery Resource Office, US Fish and Wildlife Service, Yreka, CA. Nakayama K, Oshima Y, Nagafuchi K, Hano T, Shimasaki Y, Honjo T. 2005. Early-life- stage toxicity in offspring from exposed parent medaka, Oryzias latipes, to mixtures of tributyltin and polychlorinated biphenyls. Environ Toxicol Chem 24(3):591-596. Nightingale B, Simenstad C. 2001. Dredging activities: marine issues. Prepared for Washington Department of Fish and Wildlife, Washington Department of Level II (Screening) Ecological Risk Assessment DRAFT Fred Wahl Boatyard July 2010 26

Ecology, Washington Department of Transportation. University of Washington Wetland Ecosystem Team School of Aquatic and Fishery Sciences. Nirmala K, Oshima Y, Lee R, Imada N, Honjo T, Kobayashi K. 1999. Transgenerational toxicity of tributyltin and its combined effects with polychlorinated biphenyls on reproductive processes in Japanese medaka (Oryzias latipes). Environ Toxicol Chem 18(4):717-721. NOAA. 2008. Screening quick reference tables. NOAA OR&R report 08-1 [online]. Office of Response and Restoration Division, National Oceanic and Atmospheric Administration, Seattle, WA. [Cited May 15, 2009.] Available from: http://response.restoration.noaa.gov/book_shelf/122_NEW-SQuiRTs.pdf. NOAA Fisheries. 2010. Office of protected resources: marine and anadromous fish [online]. NOAA. [Cited 6/24/10.] Available from: http://www.nmfs.noaa.gov/pr/species/fish/. ODEQ. 1998. (Updated 2001.) Guidance for ecological risk assessment: levels I, II, III, IV. Waste Management and Cleanup Division, Oregon Department of Environmental Quality, Portland, OR. ODEQ. 2001. Guidance for ecological risk assessment: level II screening update. Waste Management and Cleanup Division, Oregon Department of Environmental Quality, Portland, OR. ODEQ. 2007. Guidance for assessing bioaccumulative chemicals of concern in sediment. 07-LQ-023A. Environmental Cleanup Program, Oregon Department of Environmental Quality, Portland, OR. Riisgard HU, Larsen PS. 2005. Water pumping and analysis of flow in burrowing zoobenthos: an overview. Aqua Ecol 39:237-258. Salo EO. 1991. Life history of chum salmon (Oncorhynchus keta). In: Groot C, Margolis L, eds, Pacific salmon life histories. UBC Press, Vancouver, BC, pp 231-310. Sandercock FK. 1991. Life history of coho salmon (Oncorhynchus kisutch). In: Groot C, Margolis L, eds, Pacific salmon life histories. UBC Press, Vancouver, BC, pp 395- 46. Satoh H, Nakamura Y, Okabe S. 2007. Influences of infaunal burrows on the community structure and activity of ammonia-oxidizing bacteria in intertidal sediments. App Environ Microbiol 73(4):1341-1348. SEA. 2002. Portland Harbor remedial investigation/feasibility study sediment profile image survey of the Lower Willamette River. Prepared for Lower Willamette Group. Striplin Environmental Associates, Inc., Olympia, WA.

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Shimasaki Y, Oshima Y, Inoue S, Inoue Y, Kang IJ, Nakayama K, Imoto H, Honjo T. 2006. Effect of tributyltin on reproduction in Japanese whiting, Sillago japonica. Mar Environ Res 62:S245-S248. Shimasaki Y, Kitano T, Oshima Y, Inoue S, Imada N, Honjo T. 2003. Tributyltin causes masculinization in fish. Environ Toxicol Chem 22(1):141-144. Simenstad CA. 1983. The ecology of estuarine channels of the Pacific Northwest coast: a community profile. FWS/OBS-83/05. Division of Biological Sciences, US Fish and Wildlife Service, Washington, DC. Smith SL, MacDonald DD, Keenleyside KA, Ingersoll CG, Field LJ. 1996. A preliminary evaluation of sediment quality assessment values for freshwater ecosystems. J Great Lakes Res 22:624-638. Stern JH, Hennessy DP, Patmont CR. 2003. Improving estimates of contaminant exposure for mobile organisms: an assessment of area-weighted home range exposure estimates applied to the relationship between sediment chemistry and liver lesions in English sole. Puget Sound Research Conference 2003, Vancouver, BC. Toews DA, Brownlee MJ. 1981. A handbook for fish habitat protection on forest lands in British Columbia. Field Services Branch, Habitat Protection Division, Land Use Unit, Vancouver, BC. USFWS. 2010. Endangered species program: conserving the nature of America [online]. USFWS. [Cited 6/24/10.] Available from: http://www.fws.gov/endangered/. Veinott G, Northcote T, Rosenau M, Evan RD. 1999. Concentrations of strontium in the pectoral fin rays of the white sturgeon (Acipenser transmontanus) by laser ablation sampling- inductively coupled plasma - mass spectrometry as an indicator of marine migrations. Can J Fish Aquat Sci 56:1981-1990. Vogel S. 1994. Life in moving fluids: the physical biology of flow. Second ed. Princeton University Press, Princeton, NJ. WDFW. 2007. Letter dated March 8, 2007 to T. Do, Windward, from M. Wall, WDFW, regarding information on a tagged white sturgeon. Columbia River Sturgeon Project, Washington Department of Fish and Wildlife Region 5, Vancouver, WA. WDFW. 2010. Salmon facts: an informational guide to our state's natural treasure [online]. Washington Department of Fish and Wildlife, Olympia, WA. [Cited June 2010.] Available from:

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Wilson D. 2010. Personal communication (conversation with Helle Andersen, Windward Environmental, regarding fish species utilizing area near boatyard). Port of Toledo, Toledo, OR. June 22, 2010. Windward. 2007. Lower Duwamish Waterway remedial investigation. Baseline ecological risk assessment. Prepared for Lower Duwamish Waterway Group. Windward Environmental LLC, Seattle, WA. Windward. 2010. Level I ecological risk assessment at Fred Wahl Boatyard: memorandum to M. Camarata, Oregon Department of Environmental Quality. Windward Environmental LLC, Seattle, WA. Winsor MH, Boese BL, Lee H, Randall RC, Specht DT. 1990. Determination of the ventilation rates of interstitial and overlying water by the clam Macoma nasuta. Environ Toxicol Chem 9:209-213. Wydoski RS, Whitney RR. 2003. Inland fishes of Washington. 2nd ed. University of Washington Press, Seattle, WA.

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APPENDIX A

Appendix A. ProUCL Output Results

ProUCL Results

ProUCL 4.00.04 was used to calculate 95th upper confidence limits (UCLs) for use in the risk assessment (EPA 2009). Table 1 presents a summary of the data used to calculate UCLs.

Table 1. Summary Statistics

COPPER TBT SUMMARY STATISTICS (mg/kg dw) (μg/kg dw) Detection frequency 9/9 (100%) 3/9 (33%) Minimum detect 5.3 11 Maximum detect 92.9 110 Mean detect 39.1 61.7 Range of RLs for non-detects na 3.4 to 3.7 Mean value (calculated using half the RL for non-detects) 39.1 21.7 dw – dry weight na – not applicable RL – reporting limit TBT – tributyltin The recommended UCL for copper was 80.7 mg/kg dry weight (dw) based on the 95% approximate gamma UCL. For tributyltin (TBT), insufficient detects were available to generate a reliable UCL. As shown in Table 1, TBT was detected in only three of the nine samples, while typically at least five detected values are required before a UCL is calculated. However, the TBT data was entered into ProUCL, which generated two potential UCLs: 53 μg/kg dw based on the 95% KM (t) UCL and 110 μg/kg dw based on the 95% KM (percentile bootstrap) UCL. The full ProUCL output is provided below.

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Table 2. ProUCL Output Data – Copper

a, b, c GENERAL UCL STATISTICS FOR DATA SETS WITH NON-DETECTS

USER SELECTED OPTIONS From file WorkSheet.wst Full precision OFF Confidence coefficient 95% Number of bootstrap operations 2000

GENERAL STATISTICS Number of valid observations 9 Number of distinct observations 9 Raw statistics Log-transformed statistics Minimum 5.3 Minimum of log data 1.668 Maximum 92.9 Maximum of log data 4.532 Mean 39.13 Mean of log data 3.169 Median 22.5 SD of log data 1.14 SD 35.95 Coefficient of variation 0.919 Skewness 0.648

RELEVANT UCL STATISTICS Normal distribution test Lognormal distribution test (data not normal at 5% significance level) (data appear lognormal at 5% significance level) Shapiro Wilk Test statistic 0.822 Shapiro Wilk Test statistic 0.895 Shapiro Wilk critical value 0.829 Shapiro Wilk critical value 0.829

ASSUMING NORMAL DISTRIBUTION ASSUMING LOGNORMAL DISTRIBUTION 95% Student's-t UCL 61.42 95% H-UCL 190.9 95% UCLs (Adjusted for Skewness) 95% Chebyshev (MVUE) UCL 111.8 95% Adjusted-CLT UCL 61.61 97.5% Chebyshev (MVUE) UCL 142.4 95% Modified-t UCL 61.85 99% Chebyshev (MVUE) UCL 202.5

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GAMMA DISTRIBUTION TEST (DATA APPEAR GAMMA DISTRIBUTED AT 5% DATA DISTRIBUTION (DATA APPEAR GAMMA DISTRIBUTED AT 5% SIGNIFICANCE SIGNIFICANCE LEVEL) LEVEL) k star (bias corrected) 0.836 Nonparametric Statistics Theta star 46.83 95% CLT UCL 58.84 MLE of mean 39.13 95% jackknife UCL 61.42 MLE of SD 42.81 95% standard bootstrap UCL 57.87 Nu star 15.04 95% bootstrap-t UCL 65.11 Approximate chi square value (.05) 7.29 95% Hall's bootstrap UCL 54.24 Adjusted level of significance 0.0231 95% percentile bootstrap UCL 57.36 Adjusted chi square value 6.188 95% BCA bootstrap UCL 60.24 Anderson-Darling Test statistic 0.465 95% Chebyshev (Mean, SD) UCL 91.36 Anderson-Darling 5% critical value 0.741 97.5% Chebyshev (Mean, SD) UCL 114 Kolmogorov-Smirnov Test statistic 0.212 99% Chebyshev (Mean, SD) UCL 158.4 Kolmogorov-Smirnov 5% critical value 0.286 Assuming Gamma Distribution 95% approximate gamma UCL 80.74 95% adjusted gamma UCL 95.12 Potential UCL to Use Use 95% approximate gamma UCL 80.74 a There are only nine values in this data b It should be noted that even though bootstrap methods may be performed on this data set, the resulting calculations may not be reliable enough to draw conclusions. c The literature suggests the use of bootstrap methods on data sets with more than 10 to15 observations. BCA – bias-corrected accelerated SD – standard deviation CLT – central limit theorem UCL – upper confidence limit MLE – maximum likelihood estimate MVUE – minimum variance unbiased eliminator

Level II (Screening) Ecological Risk Assessment DRAFT Fred Wahl Boatyard Appendix A: ProUCL Results

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Table 3 ProUCL Output Data – TBT

a , b GENERAL UCL STATISTICS FOR DATA SETS WITH NON-DETECTS

USER SELECTED OPTIONS From file WorkSheet.wst Full precision OFF Confidence coefficient 95% Number of bootstrap operations 2000

GENERAL STATISTICS Number of valid data 9 Number of detected data 3 Number of distinct detected data 3 Number of non-detect data 6 Percent non-detects 66.67% Raw Statistics Log-transformed Statistics Minimum detected 11 Minimum detected 2.398 Maximum detected 110 Maximum detected 4.7 Mean of detected 61.67 Mean of detected 3.752 SD of detected 49.54 SD of detected 1.204 Minimum non-detect 3.4 Minimum non-detect 1.224 Maximum non-detect 3.7 Maximum non-detect 1.308 Number treated as non-detect 6 Number treated as detected 3 Single DL non-detect percentage 66.67%

UCL STATISTICS Normal distribution test with detected values only Lognormal distribution test with detected values only (data appear normal at 5% significance level) (data appear lognormal at 5% significance level) Shapiro Wilk Test statistic 0.998 Shapiro Wilk Test statistic 0.914 5% Shapiro Wilk critical value 0.767 5% Shapiro Wilk critical value 0.767

Level II (Screening) Ecological Risk Assessment DRAFT Fred Wahl Boatyard Appendix A: ProUCL Results

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c,d,e ASSUMING NORMAL DISTRIBUTION ASSUMING LOGNORMAL DISTRIBUTION DL/2 substitution method DL/2 substitution method Mean 21.73 Mean 1.63 SD 38.87 SD 1.702 95% DL/2 (t) UCL 45.82 95% H-Stat (DL/2) UCL 203.3 MLE Method Log ROS Method MLE yields a negative mean na Mean in log scale 0.708 SD in log scale 2.527 Mean in original scale 21.03 SD in original scale 39.28 95% percentile bootstrap UCL 44.15 95% BCA bootstrap UCL 51.42

Level II (Screening) Ecological Risk Assessment DRAFT Fred Wahl Boatyard Appendix A: ProUCL Results

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GAMMA DISTRIBUTION TEST WITH DETECTED VALUES ONLY (DATA DATA DISTRIBUTION TEST WITH DETECTED VALUES ONLY (DATA APPEAR NORMAL AT 5% NOT GAMMA DISTRIBUTED AT 5% SIGNIFICANCE LEVEL) SIGNIFICANCE LEVEL) k star (bias corrected) na Nonparametric Statistics – KM Method Theta Star na Mean 27.89 nu star na SD 33.4 A-D Test Statistic na SE of Mean 13.64 5% A-D critical value na 95% KM (t) UCL 53.25 K-S Test statistic na 95% KM (z) UCL 50.32 5% K-S critical value na 95% KM (jackknife) UCL 65.46 Assuming Gamma Distribution – gamma ROS statistics using 95% KM (bootstrap t) UCL 36.55 extrapolated data Minimum na 95% KM (BCA) UCL 110 Maximum na 95% KM (Percentile Bootstrap) UCL 110 Mean na 95% KM (Chebyshev) UCL 87.33 Median na 97.5% KM (Chebyshev) UCL 113.1 SD na 99% KM (Chebyshev) UCL 163.6 k star na Potential UCLs to use Theta star na 95% KM (t) UCL 53.25 nu star na 95% KM (percentile bootstrap) UCL 110 AppChi2 na 95% Gamma Approximate UCL na 95% Adjusted Gamma UCL na a There are only three distinct detected values in this data. The number of data may not be adequate to perform GOF tests, bootstrap, and ROS methods. b It is necessary to have four or more distinct values for bootstrap methods. However, results obtained using four to nine distinct values may not be reliable. It is recommended to have 10 to 15 or more observations for accurate and meaningful results and estimates. c DL/2 is not a recommended method. d Data have multiple DLs – use of KM Method is recommended e For all methods (except KM, DL/2), and ROS Methods), observations < largest non-detect are treated as non-detects. BCA – bias-corrected accelerated GOF – goodness of fit MLE – maximum likelihood estimate ROS – regression on order statistics DL – detection limit KM - Kaplan-Meier na – not applicable SD – standard deviation Level II (Screening) Ecological Risk Assessment DRAFT Fred Wahl Boatyard Appendix A: ProUCL Results

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REFERENCES EPA. 2009. ProUCL Version 4.00.04. Statistical software for environmental applications for data sets with and without nondetect observations [online]. Technical Support Center for Monitoring and Site Characterization, US Environmental Protection Agency. Updated February 2009. [Cited August 11, 2009.] Available from: http://www.epa.gov/nerlesd1/tsc/TSC_form.htm#instructions04.

Level II (Screening) Ecological Risk Assessment DRAFT Fred Wahl Boatyard Appendix A: ProUCL Results

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