ECOLOGICAL RISK 

ASSESSMENT

for

ATLANTIC WOOD 

INDUSTRIES 

April 1992 

Michael F. Buchman

Jay Field 

Alyce Fritz

Diane Wehner

Hazardous Materials Response Division

National Oceanic and Atmospheric 

Administration

Seattle, Washington

Robert Stuart 

Robert Dexter 

EVS Consultants, Inc., 

Seattle, Washington 

[This document has been recreated from the best available photocopy: although there 

are differences in layout from the original, the content remains essentially the same.]
UNITED STATES ENVIRONMENTAL PROTECTION AGENCY 

Region III 

841 Chestnut Building Philadelphia, Pennsylvania 19107 

April 13, 1992 

FEDERAL EXPRESS 

Atlantic Wood Industries, Inc. 

Sugar Refinery Road 

Port Wentworth, GA 31407 

Attn: Ross F. Worsham

Senior Environmental Engineer 

Re: Final Ecological Risk Assessment (ERA) for the Atlantic 

Wood Industries (AWI) Superfund Site, Portsmouth, VA 

Dear Mr. Worsham:

A final draft ERA report was provided to AWI and Keystone Environmental Resources, Inc. 

(Keystone) on October 24, 1991. We discussed this document during a December 4, 1991 telephone 

conversation and written comments from AWI were received by the U.S. Environmental Protection 

Agency (EPA) on December 19, 1991. For your information, I am enclosing the final ERA report, which 

was prepared by the National Oceanic and Atmospheric Administration (NOAA) under an Interagency

Agreement (IAG) with EPA. All comments from AWI have been incorporated into the final ERA report. 

As you know, discussions involving sediment target cleanup levels, which were developed to be 

protective of aquatic receptors, and associated ecological considerations regarding first operable unit 

remediation had been included in the final draft ERA. At my request, NOAA has removed the 

aforementioned information from this document. Target cleanup levels and related discussion are now 

included in a separate Technical Memorandum, which I will be receiving tomorrow. Upon receipt, I will 

forward a copy of this Technical Memorandum to AWI and Keystone. 

Risk to terrestrial habitats at the AWI site was also discussed in the final draft ERA. This 

discussion remains, for the most part, unchanged in the final ERA. However, in accordance with the 

scope of work outlined in the IAG, I have requested that soil target cleanup levels be developed that are 

protective of terrestrial receptors. Recommended draft target cleanup levels have been prepared and are 

currently undergoing internal review through EPA's Interagency Biological Technical Assistance Group 

(BTAG). I anticipate that BTAG will complete its review in approximately one (1) week, after which time 

I will forward a copy of a second Technical Memorandum to AWI and Keystone. 
CONTENTS

PURPOSE.................................................................................................................................. 1

SCOPE OF ASSESSMENT ............................................................................................. 1

SITE CHARACTERIZATION...................................................................................... 2

AQUATIC ECOLOGICAL RISK ASSESSMENT............................................. 2

POTENTIAL RECEPTORS............................................................................................ 2

Aquatic Habitats .................................................................................................... 5 

Habitats of Special Concern................................................................................... 6 

Species................................................................................................................... 6 

EXPOSURE ASSESSMENT ........................................................................................... 7

Contaminants of Concern ...................................................................................... 8 

Sources of Contaminants ..................................................................................... 12 

Fate and Transport Analysis................................................................................ 13 

Polycyclic Aromatic Hydrocarbons........................................................ 13 

Pentachlorophenol .................................................................................. 18 

Polychlorinated Dibenzo-p-Dioxins and Dibenzofurans........................ 18 

Trace Elements ....................................................................................... 19 

Exposure Pathways.............................................................................................. 20 

Estimated Exposure Levels.................................................................................. 20 

Uncertainty Analysis............................................................................................ 21 

TOXOCITY ASSESSMENT ......................................................................................... 21

Reported Effects of Contaminants of Concern .................................................... 21 

Polycylic Aromatic Hydrocarbons.......................................................... 21 

Pentachlorophenol .................................................................................. 26 

Polychlorinated Dibenzo-p-Dioxins and Polychlorinated 

Dibenzofurans............................................................................ 26 

Trace Elements ....................................................................................... 27 

Site-specific Toxicity Studies.............................................................................. 28 

Existing Toxicity-based Criteria and Standards................................................... 28 

 RISK CHARACTERIZATION..................................................................................... 29

Comparison Against Criteria and Standards........................................................ 29 

Comparison of Estimated Exposure Levels with Toxicity Data.......................... 30 

Adverse Biological Effects.................................................................................. 31 
Joint Action of Chemical Mixtures...................................................................... 32 

TERRESTRIAL ECOLOGICAL RISK ASSESSMENT .............................. 33

POTENTIAL RECEPTORS.......................................................................................... 33

Terrestrial Habitat................................................................................................ 33 

Habitats of Special Concern................................................................................. 33 

Species................................................................................................................. 33 

EXPOSURE ASSESSMENT ......................................................................................... 36

Contaminants of Concern. ................................................................................... 36 

Sources of Contaminants ..................................................................................... 36 

Fate and Transport Analysis................................................................................ 36 

Exposure Pathways.............................................................................................. 36 

Estimated Exposure Levels.................................................................................. 36 

Ingestion Doses....................................................................................... 36 

Dermal Exposure .................................................................................... 37 

Inhalation Exposure ................................................................................ 37 

Uncertainty Analysis............................................................................................ 37 

TOXICITY ASSESSMENT........................................................................................... 39

Reported Effects of Contaminants of Concern .................................................... 39 

Site-Specific Toxicity Tests................................................................................. 42 

Existing Toxicity-based Criteria and Standards................................................... 42 

RISK CHARACTERIZATION..................................................................................... 42

Comparison of Estimated Exposure Levels with Toxicity Data.......................... 43 

Adverse Biological Effects.................................................................................. 43 

Joint Action of Chemical Mixtures...................................................................... 43 

REFERENCES ..................................................................................................................... 44
ATLANTIC WOOD INDUSTRIES SITE ECOLOGICAL RISK ASSESSMENT 

PURPOSE 

The Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA), as amended 

by the Superfund Amendments and Reauthorization Act of 1986 (SARA), calls upon the U.S. 

Environmental Protection Agency (EPA) to protect human health and the environment with respect to

releases or potential releases of contaminants from hazardous waste sites. The National Oil and

Hazardous Substances Pollution Contingency Plan (NCP) of March 8, 1990, calls for identification and 

mitigation of the environmental impact of these sites and the selection of remedial actions that protect the 

environment. In addition, numerous federal and state laws and regulations are potentially “applicable or 

relevant and appropriate requirements” (ARARs). 

The National Oceanic and Atmospheric Administration (NOAA) has entered into an Interagency 

Agreement (IAG) with the EPA to perform an Environmental Assessment at the Atlantic Wood Industries 

Superfund Site (Figure 1). In partial fulfillment of the requirements of the IAG, NOAA has prepared an 

ecological risk assessment report. The primary objective of the ecological risk assessment is to determine

whether contaminants released to the environment at the Atlantic Wood site have impacted terrestrial and

aquatic habitats. 

SCOPE OF ASSESSMENT 

The ecological risk assessment for the Atlantic Wood site is based upon information from a variety of

sources. Information presented in the ecological risk assessment regarding the site's history, physical 

description, geology, hydrology, and analytical chemistry were taken directly from the site’s remedial

investigation (RI) report prepared by Keystone Environmental Resources (KER), Inc. (1990). Aquatic and

terrestrial resource information for the Atlantic Wood site and the surrounding vicinity was obtained 

through state and federal agencies. No site-specific information was available describing aquatic or 

terrestrial species that actually use the site or in what capacity they use it. Therefore, the resource

information is general in nature and represents species that most likely could be found at or near the site. 

The ecological risk assessment attempts to answer the following questions: 

• What organisms are potentially exposed to contaminants from the site?

• What are the significant routes of exposure? 

• To what amounts of each contaminant are organisms potentially exposed? 

• What are the potential risks to aquatic and terrestrial species that come in contact with 

contaminants at the site? 

The geographic scope of this ecological risk assessment is limited to the Atlantic Wood property; the 

associated shoreline of the Elizabeth River from the Jordan Bridge to the southern boundary of the site

and extending out into the river to a line that includes the Atlantic Wood dock; and the river inlet between

Jordan Bridge and the Atlantic Wood dock (Figure 2).

Although the geographic scope of this assessment is limited to the aforementioned areas, due to the lack

of site-specific information, ecological evaluations of the surrounding vicinity have been included to 

provide representative habitat and species information. Only the primary contaminants (based on 

concentrations, toxicity, and association with site operations) were considered in this risk assessment. Not 

all compounds detected at the site were included. Chemical or toxicological site data for some trace 

contaminants (tar acids or bases) were unavailable, and, so, were not included. 

1 
A risk characterization was developed to determine whether contaminants have impacted terrestrial and

aquatic habitats. Three approaches were used to describe the levels of soil and sediment contamination

that cause significant environmental toxicity. Information from these approaches, plus other

environmental studies can be used to develop target cleanup levels intended to protect aquatic resources. 

SITE CHARACTERIZATION 

Atlantic Wood Industries is an active wood-processing facility that has been operating since 1929. The 

site is located on the west bank of the Elizabeth River in the industrialized waterfront of Portsmouth,

Virginia. The site occupies 19.2 hectares of flat land, with elevations ranging from mean sea level to 3

meters above mean sea level (Figure 1). The eastern half of the site contains active wood-processing 

facilities and wood storage areas. The western half is used to store treated and untreated wood (Figure 2). 

Surface runoff from the site drains into the Elizabeth River via three National Pollutant Discharge

Elimination System (NPDES)-permitted stormwater outfalls plus the Elm Avenue storm sewer outfall 

(Figure 2). Runoff from the northeast portion of the site drains via the storm sewer outfall and Outfall 002 

to a small inlet of the Elizabeth River between the site and the Jordan Bridge. Outfall 001 receives runoff 

flowing east across the southeast storage area. Outfall 003, located in the northwestern corner of the 

property, discharges runoff from the western portion of the site into an open ditch that leads to Paradise 

Creek and eventually discharges to the river (Environmental Strategies Corporation (ESC), 1988). The

inlet from the Elizabeth River that runs along the northern boundary of the site also receives direct surface 

water runoff from the site. 

Two groundwater-bearing zones have been identified beneath the Atlantic Wood site, the upper Columbia 

aquifer, ranging between 5.5 and 7 meters, and the lower Yorktown-Eastover aquifer. A semi-confining 

unit of clay is located beneath the Columbia aquifer. The Columbia aquifer is considered to be a water 

table aquifer recharged predominantly from precipitation (KER, 1990). 

Within the eastern portion of the site, groundwater flows east towards the Elizabeth River. Groundwater

flow in the western portion of the site is radial due to a groundwater mound. Average linear velocity

calculated for flow beneath the eastern portion of the site is 27.5 meters per year (KER, 1990). Flow

velocity information was not available for the western portion of the site. 

AQUATIC ECOLOGICAL RISK ASSESSMENT 

POTENTIAL RECEPTORS 

Ecological receptors in three types of aquatic habitat have been identified as being at risk due to 

contaminant migration from the Atlantic Wood site. These are the estuarine communities in the Elizabeth 

River along the shoreline of the site and in the on-site inlet of the river, five very small and disturbed

wetlands onsite, and a nearby freshwater stream (Paradise Creek). Only the first two types of habitats, 

which occur within the geographical scope of this assessment, will be addressed in this risk assessment.

Organisms that are full-time residents of these habitats or that may utilize these areas during migratory

periods are considered to be potential receptors of site-related contamination. 

2 
Figure 1. Atlantic Wood Industries Site in Portsmouth, Virginia, showing sediment

sampling stations (16, 17, 18) and oyster transplant locations (12, 17, 19)

in main river (Hargis et al., 1984; Huggett et al., 1984).

3 
4 

Figure 2. Geographic scope of Ecological Risk Assessment (ESC, 1988). .
Aquatic Habitats

The Elizabeth River is the dominant aquatic habitat impacted by the site. The river system has three main

branches that empty into the southern end of Chesapeake Bay. The Atlantic Wood site is located on the 

west bank of the south branch of the river, approximately 17 kilometers (km) from the river mouth

(Figure 1). Tidal influence in the south branch extends approximately 30 km upstream from the river 

mouth, with salinities in the area of Atlantic Wood ranging from 10 to 22 parts per thousand. Intertidal 

mudflats occur along the border of the site, especially at the mouth of the inlet and Outfall 001. These 

mudflats are small and essentially devoid of macrophytes. No biological survey has been conducted to

determine the functional viability of the mudflats or river bottom bordering the site. 

According to the City of Portsmouth Tidal Marsh Inventory (July 1989), there are no remaining wetlands

along the western shoreline of the south branch of the Elizabeth River. The National Wetland Inventory 

also conducted a wetland study in this area and found no wetlands within the site boundaries. However, 

despite these earlier findings, a wetlands delineation study conducted by KER, Inc. (1990) identified five

wetlands according to the Federal Manual for Identifying and Delineating Jurisdictional Wetlands. All the

wetlands identified are in a disturbed condition and are of very low functional value based: on their small

size, low vegetation diversity, scattered vegetation, disturbed soils, and very minimal wildlife usage. 

There are basically two types of wetlands found on the site in five different locations (Figure 2). The first 

type is a reed grass community and the second a saltbush community (groundsel tree). Reedgrass 

(Phragmites australis) is an aggressive, less desirable species with little value to wildlife. It typically

invades disturbed marshes and competes with species considered more desirable for their habitat value. 

Groundsel communities are considered of moderate value for the diversity and bird nesting area they add 

to the marsh ecosystem. 

The five wetland areas can be described as: 

• Location 1 - Small, fringing, tidal zone wetland, groundsel tree community, found along the 

stormwater drainage channel entering the inlet. 

• Location 2 - Tidal zone, groundsel tree community, also small patches of saltmarsh cordgrass. 

Soil is mostly contaminated waste. 

• Location 3 - Reedgrass community in small depression along the property border receiving 

runoff from off-site areas. 

• Location 4 - Disturbed, mixed community in constructed depression formed from the removal of 

tanks from this area. Water flows continuously into this area from the plant's cooling 

tower, thus creating conditions sufficient to allow the growth of some wetland 

plants. 

• Location 5 - Reedgrass community, partially fringing the Outfall 003 drainage ditch and covering 

the previous waste fill area on the southwest corner of the site property. 

Six small tidal tributaries represent the extent of riparian habitat in off-site areas near the site, including

those creeks below the Jordon Bridge (100 meters north of the site) and the Gilmerton Bridge (3.5 km to 

the south): Paradise, Scuffletown, Jones, Gilligan, St. Julian, and Milldam creeks. Riparian habitats near 

the site are estuarine emergent marshes and stream banks. The total area of emergent marsh for these six

creeks is approximately 52 hectares, with Paradise Creek offering the largest aggregate area of

approximately 21 hectares. These wetlands are narrow fringe marshes interspersed along each tributary,

often abutting developed uplands. Vegetation is salt-tolerant grasses and sedges with low shrubs along the 

higher upland edge. Their small size limits the utility of these habitats for wildlife. According to local 

sources, these wetlands and stream banks are only considered valuable because they are scarce, not 

because they provide superior habitat (Schwab, personal communication, 1991). Overall value is 

considered to be further diminished by surface water and sediment contamination in the Elizabeth River 

5 
(Schwab, personal communication, 1991). Despite such conditions, these riparian habitats do support 

small, but stable populations of wildlife, most prominently muskrat (Halbrook, personal communication, 

1991). 

The site is located between two large riparian/estuarine wetland habitats of high value--Chesapeake Bay,

7 km to the north, and Great Dismal Swamp National Wildlife Refuge, 12 km to the south (U.S. Fish and 

Wildlife Service (USF&WS), 1980). The Elizabeth River connects these two large habitats, acting as an 

avian wildlife corridor. The site is also located under the Atlantic Flyway, the primary East Coast avian

migratory route. Therefore, in addition to supporting small resident populations of wildlife, the limited

habitats on site are also used on an intermittent, but at times moderate, basis by transient wildlife. 

Habitats of Special Concern

Based on the limited information regarding species occurrence near the site, no endangered or threatened

species are present and no critical habitats are located near the site. The closest significant habitats are 

two heron rookeries--great blue heron and yellow-crowned night heron--5 km downstream near the 

Hampton Roads Army Terminal (USF&WS, 1980). 

Species

Because comprehensive wildlife surveys at or nearby the Atlantic Wood site have been limited, 

information on the occurrence of species onsite is sporadic. No data were found regarding the presence of 

aquatic vegetation in the Elizabeth River. The river would be expected to support a normal phytoplankton

population, but macrophytes are probably limited to the shoreline wetlands. Data regarding the 

invertebrate species in the Elizabeth River were largely unavailable. The river would be expected to

support a normal estuarine population of zooplankton and benthic infauna, but the identities of such

species and their spatial and seasonal distribution have not been established. 

Blue crab (Callinectes sapidus) and eastern oyster (Crassostrea virginica), species of commercial and 

recreational fishing interest, are known to be present in the Elizabeth River (Table 1). Blue crabs are 

harvested throughout the river, both recreationally and commercially, from approximately 14.5 km

upstream of the site to the river mouth (Gillingham, personal communication, 1989; Travelstead, personal 

communication, 1989). Blue crabs are distributed differently by sex and life stage, with juveniles and 

adult males preferring less saline habitat. 

Eastern oysters are normally present throughout lower portions of the river, although the extent of their 

presence near the site has not been established. In addition, the Virginia Division of Shellfish Sanitation 

enacted a prohibition on the taking of shellfish, other than blue crabs, in the Elizabeth River in April 

1982, based on contamination by trace elements, including cadmium, arsenic, chromium, copper, lead,

and zinc. This prohibition covers the river and its tributaries from Craney Island at the river mouth, and 

includes all areas approximately 20 km upstream of the site to Oak Grove, Virginia (Wright, personal 

communication, 1991). 

The south branch of the Elizabeth River provides adult foraging habitat and juvenile nursery habitat for 

numerous anadromous, catadromous, and estuarine fish species. Species that inhabit the river near the site

and their life stages are listed in Table 1. The species in Table 1 have been observed recent samples

collected within 2 km of the Atlantic Wood site, or their range is known to include the Atlantic Wood

section of the Elizabeth River's south branch. The category of adult “foraging" habitat includes use by

resident species and migratory fish that occupy the area on a seasonal basis. Many of these species have

been found immediately offshore of the Atlantic Wood site: hogchokers (Trinectes maculatus), weakfish

(Cynoscion regalis), and oyster toadfish (Opsanus tau) are residents; while spot (Leiostomus xanthurus),

and Atlantic croaker (Micropongonias undulatus) are present during the spring and summer (Huggett et 

al., 1987; Moyle and Cech, 1982). Juveniles of these species as well as juvenile Atlantic menhaden

(Brevoortia tyrarnnus), bay anchovy (Anchoa mitchelli), alewife (Alosa pseudoharengus), and blueback 

6 
herring (Alosa aestivalis) have been sampled in the river, either directly offshore of the site or within 2

km (Virginia Institute of Marine Science (VIMS), 1989). 

Table 1. Habitat use and fisheries of the south branch of the Elizabeth River within 2 km of Atlantic 

Wood (Gillingham, personal communication, 1989; Huggett et al., 1987; Austin, personal 

communication, 1989; VIMS, 1989). 

Species Habitat Fisheries

Common Name Scientific Name Spawning Nursery Foraging Sport Comm.

Anadromous

Blueback herring Alosa aestivalis ♦ ♦ ♦ ♦

Alewife Alosa pseudoharengus ♦ ♦ ♦ ♦

American Shad Alosa sapidissima ♦ ♦ ♦

Striped bass Morone saxatilis ♦ ♦ ♦

Catadromous

American eel Anguilla rostrata ♦ ♦ ♦

Estuarine

Bay anchovy Anchoa mitchelli ♦ ♦

Blue crab Callinectes sapidus ♦ ♦ ♦ ♦

Eastern oyster* Crassostrea virginica ♦ ♦ ♦ ♦ ♦

Weakfish Cynoscion regalis ♦ ♦ ♦ ♦

Mummichog Fundulus heteroclitus ♦ ♦ ♦

Striped killifish Fundulus majalis ♦ ♦ ♦

Atlantic silverside Menidia menidia ♦ ♦ ♦

Hogchocker Trinectes maculatus ♦ ♦ ♦ ♦

Oyster toadfish Opsanus tau ♦ ♦ ♦

Marine

Atlantic menhaden Brevoortia tyrannus ♦ ♦ ♦ ♦

Spot Leiostomus xanthrus ♦ ♦ ♦ ♦

Atlantic croaker Micropongonias undulatus ♦ ♦ ♦ ♦

Spotted hake Urophycia regia ♦ ♦

* A 1982 shellfish advisory still in effect prohibits the taking of shellfish, other than blue crab, in the Elizabeth River. 

EXPOSURE ASSESSMENT 

To estimate the exposure that aquatic species have received or may receive from contaminants from the 

Atlantic Wood site requires an assessment that includes identifying each of the following: 

• which contaminants that have been detected on site are of prime ecological concern?

• what are the sources of these contaminants?

• how are the contaminants being physically transported through the environment and chemically 

altered? 

• what are the actual pathways and levels at which contaminants may migrate to habitats and

receptors of concern?

Ideally, this assessment would produce two estimates for each of the primary contaminants for potential 

receptors within the habitats of concern: 1) an estimate of the maximum possible exposure level, and 2)

an estimate of the most-likely exposure level. Finally, the assumptions and uncertainties associated with 

these estimates must be made explicit so that subsequent risk assessments reflect the appropriate level of 

probability. 

7 
8 

Contaminants of Concern

Contaminants of concern at Atlantic Wood were identified based on four criteria: prevalence, mobility, 

persistence, and toxicity. Contaminant data is from the RI (KER, 1990), which includes samples collected 

from soil, groundwater and sediment. 

The contaminants of primary concern at the Atlantic Wood site are the polycyclic aromatic hydrocarbons 

(PAHs) associated with creosote, pentachlorophenol (PCP), polychlorinated dibenzo-p-dioxins (PCDDs) 

and polychlorinated dibenzofurans (PCDFs), arsenic, copper, and zinc. Substantial data demonstrate that 

PAHs are persistent compounds that can accumulate to high levels in many invertebrate organisms and 

are toxic to most species at low concentrations. PCP and the associated chlorinated dioxins and furans 

also have been shown to be toxic to most biological species. Because creosote (PAHs), PCP, PCDDs, and 

the trace elements of concern are all known to be acutely toxic to aquatic and terrestrial species and 

because most of the toxicological information available describes acute toxicity, mortality will be the 

primary endpoint of concern for this risk assessment. Since PAHs are also known to induce a number of 

mutagenic effects, cancer, and histopathological changes in certain organisms, histopathological 

endpoints will be evaluated for the PAHs. Based on information gained from other regions regarding 

reproductive impairment due to PAH exposure, reproductive endpoints will also be considered. 

Elevated levels of PAHs and PCP were found in soil, in on-site groundwater, and in sediments of the 

drainage ditches and inlet. Elevated levels of arsenic, copper and zinc were detected in on-site soils and in 

sediments of the drainage ditches (Table 2). Additionally, a number of PCDDs and PCDFs were detected 

in soil samples (Table 2) (KER, 1990). Not all samples (e.g., inlet and river sediments) were analyzed for 

these additional analytes, and the detection limits were too high for some samples to evaluate their 

occurrence relative to toxic levels. Therefore, the distribution pattern and their risk to environmental 

receptors for PCDDs, PCDFs and trace elements cannot be discerned at this time. If information on 

additional contaminants of concern is generated, an update to this risk assessment could be conducted. 

Creosote, a wood preservative used at the site, appears to be the primary source of PAHs at the Atlantic 

Wood site. Coal tar creosote is a distillate of bituminous coal, whose properties vary depending upon the 

source of the coal tar. Creosote is typically composed of 85 to 90 percent PAHs, as well as about 10 

percent tar acids (phenols and creosols) and 5 percent tar bases (N-, S, and O-heterocyclics) (Table 3) 

(Chemical Information Systems (CIS), Inc., 1986; Mueller et al., 1989). Because PAHs represent the 

majority of creosote constituents, and because it is anticipated that the scope of remediation will be driven 

by these particular constituents, tar acids and bases will not be considered further. Although a number of 

different PAHs are present in creosote, many of the analyses performed at the Atlantic Wood site 

measured only selected PAHs. For simplicity, total PAHs (ΣPAHs) will be used here to represent the sum 

of the individual PAHs that were measured and reported. Most likely, there are other PAHs present that 

were not reported, and these would not be reflected in the ΣPAH figure. 
9 

Table 2. Concentrations (mg/l or mg/kg) of the contaminants of concern and their frequency of detection in groundw

sediments, inlet sediments and Elizabeth River sediments (KER, 1990). 

Groundwater Surface Soil Ditch Sediments Inlet Sediments 

Chemicals Mm Max Freq. Min Max Freq. Min Max Freq. Min Max Freq

PAHsAcenapthene <0.01 17000 10/30 0.05 940 74/87 <0.33 19 6/12 25 2400 5/5 Acenaphthylene <0.01 1000 S/30 0.05 42 77/87 0.47 6.3 12/12 5.7 87 5/5 Anthracene <0.01 4000 8/30 0.20 2200 80/87 1.6 560 12/12 14 3200 5/5 Benzo(a)anthracene <0.01 0.26 6/30 0.06 890 84/87 38 290 12/12 33 880 5/5 Benzo(a)pyrene <0.01 1800 S/30 0.06 840 84/87 3.4 210 12/12 27 630 5/5 Benzo(b)fluoranthene <0.01 2800 S/30 0.11 1600 84/87 5.2 240 12/12 28 1600 5/5 Benzo(g,h,i)perylene <0.01 310 S/30 0.49 300 82/87 1.8 77 12/12 9 190 5/5 Benzo(k)fluoranthene <0.01 2800 S/30 0.05 1600 85/87 62 310 12/12 24 1600 5/5 Chrysene <0.01 32000 6/30 0.07 1100 84/87 4.9 320 12/12 44 2000 5/5 Dibenz(a,h)anthracene <0.01 140 1/30 0.05 130 78/87 0.42 35 12/12 6.4 110 5/5 Fluoranthene <0.01 18000 8/30 0.17 4800 84/87 7.3 540 12/12 150 6800 5/5 Fluorene <0.01 13000 10/30 0.04 1100 73/87 0.18 37 12/12 7.9 2400 5/5 lndeno(1,2,3-c,d)pyrene <0.01 3600 3/30 0.18 380 82/87 1.8 84 12/12 9 210 5/5 Naphthalene <0.01 64000 11/30 0.10 1900 73/87 0.17 8.5 12/12 6.2 120 5/5 Phenanthrene <0.01 36000 8/30 0.04 3300 84/87 1.3 320 12/12 36 10000 5/5 Pyrene <0.01 12000 8/30 0.17 2100 83/86 7.9 480 12/12 86 5600 5/5 Total PAH <0.01 2.6 X 105 16/30 2 13254 67/87 S3 3,116 12/12 511 38437 5/5 PCP 0.004 860 10/30 0.11 970 49/85 0.42 12 10/11 <250 <250 0/5 PCDDs/PCDFs(ng/l or μg/kg)2,2,7,8-TCDD (TEF) 40.01 0.54 6/8 0.02 12.77 17/19 NA NA NA NA NA NA Trace ElementsArsenic 0.002 0.676 23/28 4 495 32/32 31.2 374 11/11 NA NA NA Copper 0.003 1.99 23/28 S 9780 32/32 95.1 1350 11/11 NA NA NA Zinc 0.007 9.5 25/28 40 20400 77/77 291 1890 11/11 NA NA NA NA: Not analyzed TEF: 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCOO) toxicity equivalency factor. 
10 

Table 3. Composition of creosote (CIS, 1986; Mueller et al., 1989). 

Chemical Percent Composition 

PAHs 85-90 

Acenaphthene 9 

Anthracaene 2 

Benzofluoranthenes 2 

Biphenyl 0.8 

Carbazole 2 

Chrysene 3 

Dibenzofuran 5 

Dimethylnaphthalenes 2 

Fluoranthane 10 

Fluorene 10 

Methylanthracenes 4 

Methylfluorenes 3 

1–Methylnaphthalene 0.9 

2–Methylnaphthalene 1.2 

Methylphenanthrenes 3 

Naphthalene 3 

Phenanthrene 21 

Pyrene 8.5 

Tar Acids 10 

Tar Bases 5 

The PAH composition of inlet sediments is similar to that found in creosote (Figure 3). To compare the 

PAH composition of inlet sediments to that of creosote, only those PAHs detected in inlet sediments that 

are also reported to be components of creosote (CIS, 1986; Mueller et al., 1989) were evaluated. The 

concentrations of the 11 selected PAHs in each sediment sample were summed and the percentage 

composition for each PAH calculated to determine the relative contribution of each PAH compound to the 

summed total. This was done for each of the four sampling locations. The range in percentage 

composition of these same selected PAHs in different creosote formulations was also calculated. In 

Figure 3, the percentage composition of each of the PAHs at each sampling location is represented by a 

vertical bar. The range of the selected PAHs in creosote is represented by the shaded area. The relative 

levels of the lighter and more volatile PAHs (naphthalene, acenaphthene, fluorene, and anthracene) are 

generally below what would be expected in creosote, but this may be due to volatilization and metabolism 

of these more labile compounds. There appears to be an increase in the relative concentrations of a 

number of the heavier and more persistent PAHs above that expected in creosote, which is most likely 

due in part, to the loss of the lighter fraction. For instance, a dense nonaqueous-phase liquid (DNAPL) 

sample consisted of 24.6, 13.8, and 6.9 percent of naphthalene, phenanthrene, and fluoranthene, 

respectively. (This sample was collected from the monitoring well nearest the inlet which happened to 

have the thickest layer of DNAPL). If it is assumed that the DNAPL is due primarily to parent creosote 

used at the site, then weathering of the creosote in soils and sediments most likely would have resulted in 

the loss of naphthalene with a resultant increase in the relative concentrations of the less volatile 

components. 
Figure 3. Average proportional PAH composition (% of ΣPAH) of sediment samples (vertical bars) from the Inlet and El

to the range of composition in different creosote formulations. 

11 
The PAH composition of sediments collected in the Elizabeth River near the site changed as the distance 

from the site increased (Figure 3). Three sediment samples were collected in a track across the Elizabeth 

River offshore from the site--at the western bank, mid-channel, and the eastern bank. The relative

concentrations from each of three sampling stations across the transect were averaged and these values 

are represented in Figure 3. Sediment samples collected on the western shore of the Elizabeth River

closest to the site had a PAH composition similar to creosote, while samples collected from mid-channel

and the eastern shore had PAH compositions showing less resemblance to creosote. The PAH 

composition of mid-channel and eastern shore sediment samples are disproportionately low in the lighter

PAHs while the heavier PAHs, particularly benzo(b)- and benzo(k)fluoranthene and pyrene are 

disproportionately high. It is possible that the mid-channel and eastern shore samples could be due to

creosote that has been in the environment longer and has thus weathered for a longer period than that in

the inlet and adjacent nearshore areas. It is also possible that there may be a different source of these

PAHs than the Atlantic Wood site. Potential sources include, or may have included, wood preserving sites 

depicted in Figure 1. 

Another widely used wood preservative and sapstain control agent used at the site is PCP. As a wood

preservative, it is carried in either a petroleum solvent or a water-soluble sodium salt. Commercial

formulations of PCP contain a variety of chemicals present as impurities (Table 4). Impurities in

commercial PCP include lower chlorinated phenols and the condensation products of two chlorinated 

phenol molecules, PCDDs and PCDFs (Firestone et al., 1972; Jensen and Renberg, 1972). Concentrations 

of the dioxin and furan impurities have been reported to range from 2 to 2,500 μg/g in technical grade 

PCP (Johnson et al., 1973). Although the most toxic dioxin congener (2,3,7,8-tetrachlorodibenzo-pdioxin) has not been detected in technical grade PCP produced in North America, the presence of other 

highly toxic PCDDs and PCDFs has been documented (Eisler, 1989). 

Table 4. Composition of commercial pentachlorophenol (Firestone et al., 1972). 

Chemical Percent Composition 

Pentachlorophenol 35-90 

Tetrachlorophenol 4-8 

Trichlorophenol <0.1 

Higher chlorophenols 24 

Caustic insolubles 1 

Sources of Contaminants

The sources of primary contaminants at the Atlantic Wood site can be associated with current or past site 

activities and the raw materials used in the treatment process. Creosote and PCP, both wood preservatives 

used by the facility, are the major raw materials from which on-site contaminants originated. A special 

formulation of creosote and PCP (creo-penta) was used from the late 1950s to the early 1960s. PCP was 

also used at the site from 1972 to 1985, and its use was briefly resumed in spring 1991. Creosote had been 

used at the site since the 1950s. However, it should be noted that all wood treatment operations were 

suspended on August 6, 1991. Although timber treated with chromated copper arsenate was, and 

continues to be, stored on the site, this compound was never employed in wood treatment operations at 

Atlantic Wood (ESC, 1988). 

Creosote was originally stored in four above-ground storage tanks located along the south side of Elm 

Avenue. Tank 1 held 3.3 million liters and the remaining three each held 1.7 million liters. Before their

removal in 1985 through 1986, these tanks contained creosote and creosote contaminated water, which 

were leaking into the storm sewer system. Since 1975, creosote being used for treatment has been stored 

in smaller tanks located in the central part of the site (ESC, 1988). 

12 
13 

Prior to 1972, the waste preservative left from the wood treatment process was stored at the southwest 

corner of the property in the “historic disposal area.” From 1972 to 1983 the “historic disposal area” was 

used to hold cuttings from the processed wood. The area was backfilled in 1983 (ESC, 1988). Additional 

information regarding past waste management practices is discussed in the RI (ESG, 1988; KER, 1992). 

Based on the results of sampling conducted during the RI, areas surrounding the treatment buildings 

contain the most heavily PAH-contaminated soils (Figure 4). Since these areas are nearest the river, it 

may represent the greatest source of contaminated runoff to wetlands of the intertidal drainage ditch, inlet, 

and river. Sampling of wetland sediments in the drainage ditch and inlet (Figure 5), and the Elizabeth 

River (Figure 6) have documented extensive PAH contamination of these areas (KER, 1990) and confirm 

transport of contaminants to these habitats. When grain size is accounted for, samples taken from the five 

stations in the inlet indicate a decreasing gradient in PAH content. The head of the inlet is dominated by 

sand and gravel with 100 percent product saturation (i.e., creosote) of the pore spaces (KER, 1990). At 

sampling sites nearer to the mouth of the inlet, the sediment texture changes to clayey sand and then 

sandy clay and there is only residual-to-heavy product saturation of pore spaces. However, these finergrained sediments allow more surface area for adsorption of PAHs, and therefore greater apparent 

concentrations than gravelly samples. 

In addition to PAHs, other contaminants were detected at the Atlantic Wood Site. Sampling across the site 

indicates widespread contamination of soils and sediments by PCP (Figures 4 and 5), although high 

detection limits associated with some samples preclude formulation of conclusions regarding relative 

source contributions to aquatic habitats. PCDDs and PCDPs were detected in soil (Figure 4) and can most 

likely be attributed to the use of PCP, which is known to contain these substances as impurities (Firestone 

et al., 1972; Jensen and Renberg, 1972; ESC, 1988). 

Elevated levels of arsenic, copper, and zinc were also detected in soils and sediments of drainage ditches 

(Figures 4 and 5). The arsenic and copper may have been derived from treated wood that was stored on 

the site. The occurrence of zinc has not been associated with the site. The occurrence of zinc has not been 

associated with current or past operational activities at the site (KER, 1992). The presence of zinc, and its 

possible association with off-site sources/activities, is currently being investigated. 

Analysis of ground water samples collected during the RI indicated the presence of PAHs and PCP in the 

surficial aquifer throughout the site (Figure 7). Moreover, PCDDs/PCDFs, arsenic, copper, and zinc were 

also detected in the surficial aquifer at elevated levels. As previously stated, DNAPL was also observed in 

several ground water monitoring wells and may serve as significant source of contamination in ground 

water. The presence of dissolved contaminants and DNAPL indicates that habitats of concern may be 

impacted. 

Fate and Transport Analysis 

Polycyclic Aromatic Hydrocarbons 

The primary source of PAHs at the Atlantic Wood site is coal tar creosote used in the wood treatment 

process. The fate and transport of creosote are largely dependent upon the physical and chemical 

properties of its components, but its behavior in the environment is complicated due to the complex 

mixture of hundreds of organic compounds. Creosote has been observed to separate into three fractions 

when mixed with seawater--floating, sinking, and dissolved (Merrill and Wade, 1985). The sinking 

fraction was reported to have a composition similar to the original sample, suggesting that this fraction 

may be transported relatively intact to sediments. Some relatively insoluble components of creosote may 

become soluble because many of the compounds occurring in creosote probably act as co-solvents to 

enhance the aqueous solubility of the less soluble component (Merrill and Wade, 1985). 
Figure 4. Maximum concentrations of total PAHs; PCP; 2,3,7,8-TCDD (TEFs); arsenic, copper, and zinc in surface soi

mg/kg, except for TCDD, which is in μg./kg. 

14 
Figure 5. Sediment sampling results at Atlantic Wood Industries (ESC 1988; Keystone Environmental Resources, 1990). 

15 
Figure 6. Sediment sampling stations in the Elizabeth River adjacent to the Atlantic Wood site (Keystone Environmental R

16 
Figure 7. Groundwater sampling results at Atlantic Wood Industries (ESC, 1988; Keystone Environmental Resources, 199

17 
Because of the relatively low aqueous solubility of PAHs and their strong affinity for particulate matter in 

aquatic systems, sediments are the primary reservoir for most of the major component creosote. Thus, the 

PAH component of creosote discharged onto surface soils and to the surface waters sorb to the soils and 

sediments near the location of discharge. Once in the sediments, creosote is subjected to burial,

resuspension, and degradation. In addition, creosote-contaminated soils and sediments can be transported 

by erosion and deposited in depositional areas away from the original source. As a result, PAHs can be

found at considerable distances from their source, as fine particles containing PAHs and other creosote 

components are transported by prevailing water currents (PTI, 1988b). Water column exposures have

been demonstrated to be significant as well (Mothershead and Hale, 1991). 

The lower molecular weight PAHs (three aromatic rings or less) predominate in the PAH assemblage of

creosote and, because of their greater aqueous solubility and degradability, will tend to be lost more

rapidly in the weathering process. Thus, in weathered creosote the less labile, higher molecular weight 

PAHs will predominate. The more polar components of creosote (i.e., the tar acids and bases) are more

soluble than PAHs of comparable molecular weight and can be expected to be preferentially lost during 

weathering. 

Although PAHs are readily absorbed by biota, they can be metabolized by the mixed-function oxidase 

(MFO) enzyme system and eliminated from most vertebrates, and, to a lesser extent, from invertebrate

species. Thus, fish exposed to PAHs would not be expected to accumulate significant levels of PAHs in

their tissues, while bivalves and some other invertebrate species could accumulate high levels (Varanasi

et al., 1985). Mothershead and Hale (1991) reported PAH bioaccumulation in newly molted blue crabs 

exposed in the vicinity of Atlantic Wood was three times higher than in intermolts. Since newly molted 

blue crab are regarded as a seafood delicacy, elevated body burdens represent a significant exposure 

pathway and may be a human health concern. Though accumulation in vertebrate organisms is usually

short term and not considered an important fate reservoir, it too can be a significant exposure pathway.

More importantly, in process of PAH metabolism, reactive intermediates are produced which are capable 

of initiating lipid peroxidation or binding to DNA. Reactions with DNA can lead to gene alterations and 

neoplasia (e.g., cancer). Thus, in converting PAHs to more water soluble forms, non-carcinogenic or procarcinogenic compounds can be converted to carcinogens. 

Microbial degradation and biotransformation are probably the ultimate fate processes for PAHs. The

available data suggest PAHs with high molecular weights are degraded and metabolized by microbes and 

multicellular organisms more slowly than lower-weight compounds. Biodegradation probably occurs 

more slowly in aquatic systems than in soil (Clement Associates, 1985). 

Pentachlorophenol 

PCP, though more soluble than PAHs, does tend to sorb to soils and sediments. In the aquatic 

environment, PCP may be present in a dissolved form, sorbed to suspended matter or bottom sediments, 

or sorbed by aquatic organisms. PCP is a neutral organic compound in aqueous solutions of pH less than 

5, but ionizes as the pH rises. The degree of dissociation will determine the solubility and degree of

adsorption to suspended matter and sediments in the aquatic system. PCP is generally considered to be 

moderately persistent in the aquatic environment (EPA, 1980). 

Uptake of PCP by organisms typically is fairly quick, but PCP can also be metabolized and excreted

quickly as well. In steady-state conditions (i.e., long-term exposures), depuration rates are usually lower

than uptake resulting in a net accumulation. A wide range of bioaccumulation values for PCP in fish and 

bivalves have been measured. 

Polychlorinated Dibenzo-p-Dioxins and Dibenzofurans 

The PCDDs and PCDFs have very similar structures and are composed of two-ring structures linked by 

oxygen. There are 135 possible isomers of PCDDs and PCDFs each, depending on chlorine substitutions 

on the rings. Because of their high octanol-water partition coefficients, PCDDs and PCDFs tend to be 

18 
highly sorbed to particulates, sediments, and biota. Their low vapor pressures mean they are unlikely to 

volatilize in the atmosphere. PCDDs and PCDFs do not readily undergo photodegradation unless solvents 

are present that will act as hydrogen donors during reductive dechlorination. Due to their tendency to be 

sorbed to particulate material and their environmental persistence, PCDDs and PCDFs can be transported 

through wind erosion or through surface water runoff. In aquatic environments, these substances can be

transported to depositional areas by suspended particulate matter. 

Some microbial degradation of PCDDs has been reported, but it appears to be limited and highly

dependent on the isomers present The half-life in soil has been found to range from 130 days to several 

years (Eisler, 1986). Uptake by biota is rapid, and bioaccumulation from PCDD-contaminated freshwater

and marine sediments has been well documented (Kuehl et al., 1986; Kuehl et al., 1987; EPA,1990; 

Rubenstein et al., 1990). The ability to metabolize PCDDs and PCDFs with low chlorine substitution via 

MFO activity has been demonstrated in some species of fish and birds but not in bivalves (Konasewich et

al., 1982). Sedimentation and bioaccumulation seem to be the most significant environmental fate for

these compounds. 

Trace Elements 

Arsenic has four different oxidation states, and chemical speciation is important in determining arsenic’s 

distribution and mobility. The +3 and +5 valence states, as well as organic complexation, are the most

important forms of arsenic in the environment. Arsenic is generally quite mobile and sorption to sediment

can be an important fate process. In the aquatic environment volatilization is important when biological

activity or highly reducing conditions produce arsine or methylarsenics (Clement Associates, 1985). 

Copper has two oxidation states, +1 (cuprous) and +2 (cupric). Cuprous copper is unstable in aerated water 

over the pH range of most natural waters (6 to 8 standard units) and oxidizes to the cupric state. Several

processes determine the fate of copper in the aquatic environment: formation of complexes, especially

with humic substances; sorption to hydrous metal oxides, clays, and organic materials; and 

bioaccumulation. In waters with soluble organic material, copper can form complexes with organic

ligands, thus favoring prolonged dispersion of copper in solution. The presence of organic acids can also 

lead to the mobilization of copper from the sediments to solution. Sorption processes are quite efficient in

scavenging dissolved copper and in controlling its mobility in natural unpolluted waters. The levels of 

copper able to remain in solution are directly dependent on water chemistry. Generally, ionic copper is 

more soluble in low pH waters and less soluble in high pH waters (Clement Associates, 1985). 

As an essential nutrient, copper is accumulated by plants and animals, although apparently it is not 

generally biomagnified. Because copper is strongly bioaccumulated and because biogenic ligands play an 

important role in completing copper, biological activity is a major factor in determining the distribution 

and occurrence of copper in the ecosystem (Clement Associates, 1985). 

Because many copper compounds and complexes are readily soluble, copper is among the more mobile of 

trace elements. The major process that limits the environmental mobility of copper is adsorption to

organic matter and clays (Clement Associates, 1985). 

Zinc can occur in both suspended particulates and dissolved forms. Dissolved zinc may occur as the free 

(hydrated) zinc ion or as dissolved complexes and compounds with varying degrees of stability and 

toxicity. The predominant fate of zinc in aerobic aquatic systems is sorption of the divalent cation by

hydrous iron and manganese oxides, clay minerals, and organic material. Concentrations of zinc in 

suspended and bed sediments always exceed concentrations in ambient water. Zinc tends to be more

readily sorbed at higher pH than lower pH, and tends to be desorbed from sediments as salinity increases

(Clement Associates, 1985). 

Atmospheric transport of zinc is also possible. However, except near major sources such as smelters, zinc

concentrations in air are relatively low and fairly constant (Clement Associates 1985). 

19 
Because it is an essential nutrient, zinc is strongly bioaccumulated even in the absence of abnormally high 

ambient concentrations. Zinc does not appear to biomagnify through the food web however. Although 

zinc is actively bioaccumulated in aquatic systems, biota appear to represent a relatively minor sink 

compared to the sediments (Clement Associates, 1985). 

Exposure Pathways

Environmental sampling at the site and in the Elizabeth River confirms significant pathways and

migration of PAHs and PCP to sediments in the wetlands of the intertidal drainage ditch, inlet, and the 

nearshore areas along the river bank (Figures 4 and 5). Besides confirming pathways, sediment residue 

data indicate an extreme level of contamination within these habitats. Sediments are thus a primary source 

of contamination to aquatic resources both directly and indirectly through desorption to overlying water 

column. Aquatic species can be exposed to contamination from the Atlantic Wood site through contact 

with contaminated sediments during residence or temporary visits to the inlet and nearshore areas of the

site. Species residing in or on the sediments, or feeding on benthic or infaunal organisms, are expected to 

receive the greatest exposure contamination from the site. In addition to ingestion or contact with 

contaminated sediment by demersal species, contaminated sediments represent a significant source and 

pathway to pelagic aquatic species using wetlands of the intertidal drainage ditch, the other wetlands on 

site, or the river inlet. These organisms may be exposed to aqueous contaminants that have leached from 

contaminated sediment, or come from outfall effluents and surface runoff. 

Estimated Exposure Levels

Estimates of exposures are based on existing knowledge of the levels of contamination across the site

derived from the RI sampling. The contaminants of concern, their concentrations, and frequency of

detection are summarized by media type in Table 2. Figures 3 through 6 show some of the groundwater,

surface soil, and sediment sampling locations with the highest concentrations of the contaminants of 

concern. 

Two estimates for sediment exposure levels of PAHs will be used for assessing risk to aquatic species. 

The maximum value detected in sediments of the inlet, on-site wetlands, or Elizabeth River adjacent to

Atlantic Wood, will be used as the maximum sediment exposure level. This value of 38,437 mg/kg ΣPAH

occurs near the mouth of the inlet and will be used to indicate a “worst-case scenario.” Because organisms 

would not be continuously exposed to this maximum possible level, it is recognized mat this maximum is 

not likely to be the long-term exposure concentration for resident organisms. Therefore, the median of 

these samples, 1,008 mg/kg, will be used as a “more-likely” sediment exposure level. This median was 

derived from 18 values that include the samples shown in Figure 5, the sample nearest Atlantic Wood 

shoreline in Figure 6 (1,059 mg/kg), and a surficial sample of Huggett (1985) collected adjacent to the 

Atlantic Wood site (1,100 mg/kg). These data points, though from different sources, were obtained by 

similar analytical techniques and all represent observations of contaminant levels in the nearshore

sediments of the Elizabeth River within the geographic scope of this assessment. 

Although PCP was reported in soil samples collected near the Elizabeth River at concentrations as high as 

970 mg/kg and in wetland sediments of an on-site drainage ditch at a concentration of 12 mg/kg, no PCP

was reported in sediments collected from the inlet or the Elizabeth River adjacent to the site (KER, 1990).

However, it should be noted that the detection limits for PCP in inlet sediments and sediments of the 

Elizabeth River (250 and 8.9 mg/kg, respectively) were relatively high. Given such elevated detection 

limits, it is quite possible that PCP could be present at significant concentrations in the sediments of these 

areas. In the absence of riverine sediment values, concentrations in soil samples will be used as 

surrogates. The maximum level of PCP in soils near the inlet, 67 mg/kg, will be used as the “high” 

exposure level to determine risk potential. The median for all soil samples, 15 mg/kg, will be used as the 

“more-likely” exposure level. 

Since metals were not measured in all samples, estimates of sediment exposure levels for metals will be 

20 
based on ten samples from drainage ditches. The median and maximum, respectively, are 135 and 364 

mg/kg for arsenic; 376 and 1,350 mg/kg for copper; and 729 and 1,890 mg/kg for zinc. 

Uncertainty Analysis

In performing the aquatic exposure assessment for the Atlantic Wood site, uncertainties arose from two 

main sources. First, those uncertainties associated with the distribution, transport, and fate of compounds 

in the environment. Second, those uncertainties associated in the estimation of chemical intakes resulting

from contact by a receptor with a particular medium (KER, 1990). 

As discussed in the RI for the Atlantic Wood site (KER, 1990), samples were not collected randomly

throughout the site, but were instead focused on areas suspected to have elevated levels of site 

constituents. Therefore, the resulting soil and sediment data were biased toward elevated concentrations. 

Thus, exposures estimated from these data would also be biased because potentially exposed species 

would most likely not spend the majority of their time in just those regions with the highest soil or 

sediment concentrations. 

To estimate an intake, certain assumptions must be made about exposure events, exposure duration, and 

the corresponding assimilation of constituents by the receptor. Because very little is known about which 

species use the site, in what capacity, and to what extent, there is a great degree of uncertainty associated

with the estimated doses. In the extreme dose scenario, it is highly unlikely that most species utilizing the

site would feed exclusively on the site and only in those areas where they would be exposed to the highest

contaminant levels. 

TOXICITY ASSESSMENT 

This section summarizes what is known about the sediment toxicity of the contaminants of concern at the 

Atlantic Wood site to aquatic species. Information on toxicity has been derived from literature reports of

bioassays with the chemicals of concern, from regional ambient toxicity surveys, and by evaluation of 

toxicity-based criteria and standards. Although this assessment of toxicity is limited to the contaminants 

of concern through sediment exposure, it is recognized that the toxicity of additional contaminants 

present at the site (e.g., tar acids and bases) may also impart significant adverse effects. 

Reported Effects of Contaminants of Concern 

Polycyclic Aromatic Hydrocarbons 

PAHs vary substantially in their toxicity to, and bioaccumulation by, aquatic species. PAHs can have

adverse impacts on all biological levels of measure--from enzymatic and immunological changes, through

tissue damage, to mortality and community alteration. The potential impact of PAHs will be discussed for

all of these levels of biological organization, with particular emphasis on impacts or species reported from

within the site’s region. 

Bioaccumulation 

In most fish, PAHs are rapidly metabolized and excreted so that levels in edible tissue are generally low

(Varanasi et al., 1989). Invertebrates, especially mollusks, do not metabolize PAHs as efficiently and may

accumulate high concentrations of PAHs in tissue or may retain detectable levels for a period of days to 

weeks (Mix, 1979; Eisler, 1987). Presence of PAHs in such species is a clear indication of bioavailability

and recent exposure. 

Oysters transplanted from relatively “clean” areas to locations throughout the Elizabeth River were used

to determine the bioavailability of PAHs. Huggett and co-workers found that the uptake rate of PAHs was 

highest at the station nearest the Atlantic Wood facility (Huggett et al., 1987) where sediment ΣPAHs

were approximately 30 mg/kg. Tissue residues of ΣPAH were also greatest at the sites off Atlantic Wood 

21 
and just downstream, and reached levels twice as high as any other stations. Except for a station near a

historical spill of pure creosote, sediment concentrations of PAHs were the highest in sediments off 

Atlantic Wood or just downstream of the site. Elevated levels of PAHs in the tissues of oysters have also

been associated with reduced condition index1

. 

PAHs are metabolized and transformed by a family of MFO enzymes that includes the enzymes

cytochrome P-450, P-450E, and ethoxyresorufin O-deethylase (EROD). These proteins are sensitive to 

environmental levels of xenobiotics and have been established as reliable indicators of exposure to

xenobiotics, including PAHs, in feral fish populations. One study of spot collected from the Elizabeth

River found levels of total cytochrome P-450, P-450E, and EROD elevated in intestine and liver 

microsomes (Van Veld et al., 1990). Fish were collected over a PAH concentration gradient in the

sediments that ranged from 0.009 to 96 mg/kg dry weight (see Figure 8 for sampling locations). Intestinal

P-450E was near the lower limits of detection in fish collected at the relatively clean sites but was 

elevated 80- to 100-fold in fish collected from contaminated sites. Intestinal EROD activity exhibited a 

similar trend. Liver P-450E and associated EROD activity was detectable in all samples and was induced 

approximately 8-fold at the most heavily contaminated sites (Table 5). 

Immune 

Other studies report additional sublethal and subcellular effects due to PAH exposure. Hogchoker from 

the Elizabeth River, where sediments contained approximately 100 rag/kg ΣPAH, were found to have

reduced macrophage activities (Weeks and Warinner, 1984). Further studies indicated that the 

macrophages had diminished chemotactic capability (the ability to migrate toward a stimulus). This 

depression on the cellular immune system could result in increased susceptibility to disease (Weeks and

Warinner, 1986; Weeks et al., 1986). Studies of oyster hemocytes, which are believed to play an

important role in defense against pathogens, have also shown depressions on the cellular immune 

systems. 

Another study of PAH effects on immune systems may add to an understanding of tumor development in

fish due to PAH exposure. Faisal et al. (in press) reported significant alterations of the in vitro tumorilytic 

activity of kidney and spleen leukocytes that defend against cancer. Mummichog (Fundulus heteroclitus) 

were collected from two sites heavily contaminated with PAHs in the Elizabeth River, plus a reference 

site. One class of cells that should spontaneously recognize their targets was unable to do so in fish

collected from areas of sediment containing 2,200 and 61 mg/kg ΣPAH. A second class of defense cells

that require prior exposure to the tumor cell antigen was found to be active, indicating an ongoing

oncogenic process in these fish. Fish from the reference site with 3 mg/kg total sediment PAHs, displayed 

neither of the oncogenic indications observed in the Elizabeth River. 

 1 Reduced condition index it a measure of tissue health as indicated by lipid content.

22 
Table 5. Total sediment PAH levels (mg/kg) and MFO activity in spot intestine and liver in fish collected

from the lower Chesapeake Bay and Elizabeth River (Van Veld et al., 1990). Refer to Figure 8 

for station locations. 

Station

Parameter 1 2 3 4 5a 6 

Total PAH1 0.009 0.055 3.1 16.0 96.0 43.0 

EROD2

Gut 47±20* 105±58* 915±320** 1550±1068** 985±319* ** 1053±594** 

Liver 230±100 267±52 652±207** 981±678** 1894±813** 1291±682** 

P450E3

Gut <0.001* 0.015±0.001* 0.079 ±0.004* 0.0109 0.131 ±0.880* 0.088 ±0.041*

Liver 0.012±0.003 0.014±0.003 0.042 0.042 ±0.004** 0.280 0.200 

Cyt. P-4504

Gut 0.43±0.05* 0.47±0.15* 0.67 0.76±0.22* ** 0.57±0.18* 0.53±0.15* 

Liver 0.78±0.15 1.01±0.18 1.00±0.28 1.14±0.31** 1.45±0.25** 1.14±0.09** 

a: Adjacent to the Atlantic Wood site 

1: Individual PAHs were not identified 

2: Values are pmoles resorufin produced per mg of microsomal protein (mean ± SD; N=7) 

3: Values are nmoles of P-450E equivalents per mg of microsomal protein (mean ± SD; N=3) 

4: Values are nmoles of cytochrome P-450 per mg of microsomal protein (mean ± SD; N=7) 

*: Significantly different from corresponding liver value P<0.05 

**: Significantly different from value determined for reference site (station 1) P<0.05

Figure 8. Sampling sites for total sediment PAH levels (rag/kg) and MFO activity in spot collected from 

the lower Chesapeake Bay and Elizabeth River (Van Veld et al., 1990). 

23 
Tumors and other tissue abnormalities have been observed in resident Elizabeth River fish. Abnormalities 

observed included fin erosion in hogchoker (Trinectes maculatus) and toadfish, plus cataracts in spot, 

weakfish, and croaker (Table 6; Huggett et al., 1987). Prevalences of these conditions increased at

stations with sediment PAHs greater than approximately 16 mg/kg. High prevalences of idiopathic

hepatic lesions have also been found in mummichog collected in the Elizabeth River adjacent to the 

Atlantic Wood site. Grossly visible hepatic lesions occurred in 93 percent of the fish from this location

where sediment contained 2,200 mg/kg ΣPAH. Hepatic lesions were not detected in fish from two less 

contaminated sites where total sediment PAHs were 3 and 61 mg/kg (Table 7; Vogelbein et al., 1990). 

The association between PAHs and liver lesions has been made in other regions of the country as well. 

For example, studies of English sole (Parophrys vetulus) from a creosote-contaminated harbor in Puget

Sound, Washington indicate statistically significant correlations between the sediment concentrations of

PAHs and the prevalence of hepatic neoplasms and other liver lesions. Similar correlations are also shown

between the concentrations of PAH metabolites in the bile (Malins et al., 1987). In addition, high levels

of PAHs in stomach contents suggest that the consumption of contaminated benthic invertebrates is an

important exposure pathway (Malins et al., 1985). 

Table 6. Total sediment PAH concentrations (mg/kg dry weight), biomass, number, and percentage of 

fish showing gross abnormalities at various sites along the Elizabeth River (Huggett et al., 

1987). 

River Kilometers from Atlantic Wood Site 

Downstream Adjacent Up-stream

Parameter 

10.5 8.5 6.5 4.5 2 0 -2

Total PAH1 1.10 5.00 5.10 11.50 15.82 24.4 154.4 

Total number of fish 1,335 1,200 1,125 705 380 320 245 

Biomass (kg) 34 26 21 13 9 7 4 

Fin Erosion (% incidence)

Hogchoker 0.7 0 0 0.4 1.4 5.5 4.3 

Toadfish 0 0 11.0 5 0 11.5 30.1 

Cataracts (% incidence)

Spot 0 0 0.1 0 3.0 0.8 9.6 

Weakfish 0.2 0 0 0.8 1.0 1.8 3.5 

Croaker 3.3 1.4 1.5 2.2 4.5 7.9 15.8 1

 Total PAH is the sum of the PAHs listed by Huggett et al. which were common to those measured at the Atlantic Wood site 

and includes phenanthrene, fluoranthene, pyrene, benzo(a)pyrene, chrysene, the benzo(a)- and benzo(b)fluoranthene, 

benz(a)anthracene, indeno(c,d-1,2,3)pyrene, and benzo(g,h,i)perylene. 

24 
Table 7. Total sediment PAH concentrations and lesion prevalences for mummichog from three 

sampling locations in Chesapeake Bay (Vogelbein et al., 1990). 

Percentage of Fish with Lesions 

Lesion Types Ware River (N=15) Scuffletown Creek 

(N=30)* 

Elizabeth River 

(N=60)** 

Total sediment PAHs (mg/kg dry wt)

1 3 61 2200

Foci of cellular alteration 0 0 73.3 

Cholangiolar proliferation 0 0 11.7 

Hepatocellular adenoma 0 0 1.77 

Early hepatocellular carcinoma 0 0 13.3 

Advanced hepatocellular carcinoma 0 0 20.0 

Total hepatocellular neoplasms 0 0 35.0 

* : Tributary to the Elizabeth River directly across from the Atlantic Wood site 

**: Adjacent to the Atlantic Wood site 1

 Total PAH as measured by Vogelbein et al. Individual PAHs were not identified. 

Studies from other regions have also documented the potential for PAH exposure to disrupt the

reproductive cycle of fish. Another study of English sole from Eagle Harbor, Washington provided

evidence of reduced reproductive success in association with high levels of sediment PAHs (Johnson et

al., 1988). Females from highly contaminated sites were found to be less likely to undergo gonadal 

maturation and to have lower mean levels of plasma estradiol (a steroid sex hormone) than those from 

cleaner sites. Both these conditions were highly correlated with aryl hydrocarbon hydroxylase (AHH)

activity, and were also in concordance with general levels of PAH metabolites in the bile. Similar findings 

for starry flounder (Platichthys stellatus) from San Francisco Bay, California have been reported. Females 

collected from contaminated sites exhibited higher AHH activity, lower levels of plasma steroids, lower 

levels of fertilization success, and decreased normal embryologic developmental success (Long and 

Buchman, 1989). 

Mortality 

At greater concentrations, PAHs are also capable of causing acute mortality. One laboratory study in

which the freshwater amphipod, Diporeia sp., was exposed to sediments dosed with a PAH mixture 

reported mortality at ΣPAH concentrations in the range of 100 mg/kg (dry weight) after 26 days of 

exposure (Landrum et al., 1991). The rate of PAH accumulation was also examined and found to be

dependent upon the sediment PAH concentration. At high exposures, uptake was greater than that 

predicted through measured partitioning between interstitial water and sediment. 

Alden and Butt (1987) conducted sediment toxicity bioassays on sediments from 16 stations in the 

Elizabeth River using the grass shrimp Palaemonetes pugio. Few biological effects were observed in

sediments with PAH concentrations in the range of 0.081 to 11 mg/kg. However, sediments from the 

south branch of the Elizabeth River near the Atlantic Wood site with PAH concentrations in the range of 

5.8 to 65.5 mg/kg were reported to produce rapidly lethal and sublethal effects. 

Bioassays exposing juvenile spot to creosote-contaminated sediments from the south branch of the 

Elizabeth River (near Republic Creosote Co.) demonstrate that contact with sediment and/or the water

associated with these contaminated sediments may result in acute death or serious sublethal effects. Skin 

lesions and mortality were observed in some of the test fish after only 8 days of exposure to sediment 

containing approximately 2,500 mg/kg ΣPAH. After 28 days, there was over 50 percent mortality. All the 

test fish exposed to the contaminated sediments showed sublethal abnormalities (skin lesions, fin erosion, 

gill erosion, reduced hematocrit level, pancreatic and liver alterations, and lack of weight gain). These 

same abnormalities were observed in several to all test fish that were exposed only to the effluent from 

25 
the test tank containing the contaminated sediment (Hargis et al., 1984; Roberts et al., 1989). Some of the 

same adverse effects have been observed in feral fish collected from the Elizabeth River near the site. 

If PAH exposure imparts acute mortality to individual organisms, a cumulative impact on overall 

community structure would be also expected. A study of benthic macroinvertebrate assemblages in 

Elizabeth River sediments concluded that the vertical distribution of organisms was affected in the areas

of highest PAH contamination. In areas of high hydrocarbon contamination, few organisms were found

below a depth of 5 cm (Diaz, 1985). Other field experiments testing the colonization of creosotecontaminated sediment by benthic invertebrates indicated that polychaetes and crustaceans were

significantly affected at creosote concentrations greater than 177 mg/kg. Mollusks, however, were

apparently more resistant to the effects of creosote contamination, showing effects only at concentrations 

greater than 844 mg/kg (Tagatz et al., 1983). 

Fish distributions along the Elizabeth River have also exhibited adverse community impacts. Huggett et 

al. (1987) reported reduced total biomass, reduced total numbers of individuals, and reduced abundance

of selected species (Table 6). These factors, as well as the increased prevalence of several gross

abnormalities noted above, correlated with PAH contamination in the sediments and proximity to the

wood treatment facilities. The largest shift in fish assemblages seem to occur when sediment PAH 

concentrations were greater than approximately 12 mg/kg. 

Pentachlorophenol 

No studies were found relating sediment concentrations of PCP to toxicity in aquatic organisms. 

However, a number of studies have established the aqueous toxicity of PCP to both freshwater and marine 

species (Choudhury et al., 1986). PCP affects energy metabolism and increases oxygen consumption. 

Collectively, its effects would lead to reduced availability of energy for growth, reproduction and reduced 

ability to avoid predators or to catch prey. Acute toxicity values (LC50) for sensitive marine and 

freshwater fish species are less than 100 μg/1. Abnormal development of oyster larvae was reported at

levels of 40 mg/l, with acute toxicity at 77 mg/1 (USF&WS, 1989). Reproductive impacts have been 

reported for a variety of sensitive species at levels ranging from 15 to 100 mg/1 (USF&WS, 1989).

Results from a study with an experimental ecosystem showed that low levels of PCP (15.8 μg/1) led to a 

reduction in numbers of individuals and species in a marine benthic community, with mollusks showing 

the most sensitivity (Tagatz et al., 1978). 

Fish may bioconcentrate PCP to levels 10,000 times the concentration in water (Eisler, 1989). 

Measurements of PCP levels in six species of fish representing two trophic levels from Lake Ontario

suggested that PCP accumulation in fish is primarily through direct uptake from water rather than 

accumulation through the food chain (Niimi and Choy, 1983). 

Polychlorinated Dibenzo-p-Dioxins and Polychlorinated Dibenzofurans 

PCDDs or PCDFs were not reported in drainage ditch sediments or sediments of the inlet. However, due

to high detection limits, the presence and spatial occurrence of the chemicals cannot be evaluated. Both

compounds were detected in on-site soil and groundwater and could be of concern in nearshore habitats. 

Dioxins and dibenzofurans, particularly 2,3,7,8-TCDD and 2,3,7,8-TCDF, are known to be toxic to 

aquatic organisms at extremely low concentrations. Very little information is available on the toxicity of

other dioxin congeners to aquatic organisms, but some are also known to be toxic at very low levels.

Dioxins and furans are environmentally stable and have a strong tendency to accumulate in lipid tissues of 

aquatic organisms. Bioavailability of 2,3,7,8-TCDD in freshwater and marine sediments to aquatic 

organisms has been demonstrated (Kuehl et al., 1986; Rubenstein et al., 1990). Effects associated with

dioxins and furans include lethality, carcinogenicity, mutagenicity, teratogenicity, reproductive

impairment, immunotoxicological, and histopathological abnormalities (Eisler, 1986). Toxicity is related

to duration of exposure as well as concentration (EPA, 1990). The toxic effects following even short-term 

TCDD exposure are commonly delayed 30 to 80 days. Fish species appear to be more sensitive than other 

26 
aquatic organisms. Developing embryos are the most sensitive life stage to dioxin effects (Cooper, 1989;

Spitzbergen et al., 1991; Walker et al., 1991). 

The U.S. Food and Drug Administration (FDA) has issued a health advisory for concentrations of 2,3,7,8-

TCDD over 50 pg/g (parts per trillion) and suggested that concentrations less than 25 pg/g in fish tissue 

should not cause serious human health risk (EPA, 1984). Ambient water quality criteria (AWQC) for the 

protection of aquatic organisms are available for freshwater only (EPA 1984). Acute AWQC was not 

determined due to the lack of sufficient toxicity data on TCDD. The “chronic” AWQC value (final

residue value) of 10 picograms per liter for 2,3,7,8-TCDD, based on a bioconcentration factor of 5,000, is

the concentration estimated to result in average levels in edible fish tissue that exceed the PDA advisory

levels (EPA, 1984). However, a recent EPA report concluded that levels of 0.038 pg/1 for 2,3,7,8-TCDD 

(or TCDD-equivalents) and 0.41 pg/1 for 2,3,7,8-TCDF would be expected to result in toxicity in some 

aquatic species (EPA, 1990). 

Trace Elements 

The toxicity of arsenicals to aquatic life is modified by numerous biological and abiotic factors. Arsenic is

carcinogenic and teratogenic in humans and other mammals. Both acute mortality and sublethal effects 

have been observed in fish and invertebrates. Acute toxicity has been shown to be affected by water

temperature, pH, Eh, organic content, phosphate concentration, suspended solids, and presence of other

substances and toxicants, as well as arsenic speciation and duration of exposure (Eisler, 1988). Early 

developmental stages are the most sensitive to arsenic toxicity. Some aquatic species have been reported 

to be adversely affected at water concentrations of 19 to 48 μg/1, or 120 mg arsenic/kg in the diet, or 

tissue residues of 1.3 to 5 mg arsenic/kg fresh weight (Eisler, 1988). 

Arsenic has been shown to bioconcentrate in organisms but is not biomagnified in the food chain.

Bioconcentration factors for inorganic arsenic in most aquatic invertebrates and fish exposed for 21 to 30 

days did not exceed 17 (Eisler, 1988). 

Information on the toxicity of sediment-associated arsenic is scarce (Long and Morgan, 1990). Data from 

Puget Sound, Washington indicate that arsenic levels of 63 mg/kg are moderately toxic to amphipods: 

nearly one-fourth of the sampled oyster larvae developed abnormally following exposure to sediment 

containing levels of 59 mg/kg. Benthic community structure was altered when sediment concentrations 

were above 57 mg/kg. Unfortunately, no spiked sediment bioassays or laboratory studies were available 

to confirm impact levels suggested by these field observations. 

The toxicity of copper to aquatic life has been shown to be related primarily to activity of the cupric (+2) 

ion, and possibly to some hydroxy complexes (EPA, 1985). The cupric ion is highly reactive and forms 

moderate to strong complexes and precipitates with many inorganic and organic constituents of natural

waters. It is readily sorbed onto surfaces of suspended solids. The proportion of copper present as the free 

cupric ion is generally low and may be less than one percent in waters where complexation predominates.

Most organic and inorganic copper complexes and precipitates appear to be much less toxic than free 

cupric ion and tend to reduce toxicity attributable to total copper (EPA, 1985). 

As an essential nutrient, copper is accumulated by plants and animals, although it is not generally 

biomagnified. Among saltwater species, the highest bioaccumulation factors are those for the bivalve 

mollusks. Oysters can bioaccumulate copper up to 28,200 times ambient levels without significant 

mortality (Clement Associates, 1985). 

Several observations on sediment toxicity of copper are available. Phelps et al. (1983) reported that 

burrowing time in littleneck clams, Protothaca staminea, was increased when copper concentrations in 

marine sediments were above 17.8 mg/kg dry sediment. Burrowing times of Macoma were significantly 

longer at concentrations of 67 mg/kg (McGreer, 1979). Levels above 19.5 mg/kg were found to be toxic 

to Hyalella (in Long and Morgan, 1990), while significant mortality to Daphnia has been reported at 

levels of 68 mg/kg (Qasim et al., 1980). Significant mortality to amphipods and abnormal development of 

27 
bivalve larvae have occurred in San Francisco sediments with copper levels above 64 and 76 mg/kg, 

respectively.

Zinc may be found in several chemical forms in natural waters. The dominant fate for zinc is sorption to

sediments, but this process is influenced by pH, salinity, and the presence of organic complexes. Reported

acute toxicity values for saltwater fish range from 2,730 to 83,000 μg/l and from 166 to 55,000 μg/l for 

saltwater invertebrates. Zinc produces chronic toxicity in the mysid shrimp (Mysidopsis sp.) at 166 μg/l. 

Adult mummichog exposed to 43,000 μg/l of zinc for 192 hours showed no adverse effects; whereas, a 

concentration of 157,000 μg/l resulted in 100 percent mortality (EPA, 1987). 

Zinc is another essential nutrient and is bioregulated by most species. Bioconcentration factors in edible

portions of aquatic organisms range from 43 for the soft-shell clam to 16,700 for the oyster (Clement 

Associates, 1985). 

In a spiked-sediment bioassay with the marine amphipod Rhepoxynius abronius, the 10-day LC50 for zinc 

was 276 mg/kg (Swartz et al., 1988). In Puget Sound, field-collected sediments were always toxic to R. 

abronius at zinc concentrations greater than 960 mg/kg (n=287) and resulted in a significant decrease in 

the abundance of one or more major macroinvertebrate taxa (polychaetes, mollusks, crustaceans, or total 

benthic infauna) at concentrations greater than 410 mg/kg (n=20). 

Site-Specific Toxicity Studies

The only direct measure of toxicity of site-specific contaminants migrating from the Atlantic Wood site 

are the result of biomonitoring requirements for Atlantic Wood's NPDES permit. The NPDES permit 

specifies that 48-hour static acute bioassays be performed annually with effluent samples from outfalls 

#001, #002, and #003 to determine the acute mortality due to exposure to outfall effluents. Bioassays 

conducted on effluent samples collected in 1986 from two of the Atlantic Wood stormwater outfalls 

showed acute toxicity to mysid shrimp Mysidopsis bahia at effluent concentrations of 47 and 37 percent

(i.e., 37 % effluent and 64% diluent water) for Outfalls 001 and 002, respectively (ESC, 1988). No 

toxicity was observed in tests with sheepshead minnows (Cyprinodon variegatus) and mysid shrimp

exposed to 1991 samples (Zieske, 1991). Data for other years (1984 to 1989) were not provided. 

Existing Toxicity-Based Criteria and Standards

The only criteria and standards applicable to the Atlantic Wood site are the EPA AWQC for the 

protection of aquatic organisms and the site-specific NPDES effluent permits. Applicable AWQC for 

marine surface waters are presented in Table 8. 

Table 8. AWQC for the protection of marine organisms for the chemicals of concern at the Atlantic

Wood site (EPA, 1986g). 

Marine AWQC (μg/l) Chemicals of Concern Acute Chronic 

PAHs

Acenaphthene 970* 710* 

Fluoranthane 40* 16* 

Naphthalene 2,350* NA 

PAH 300* NA 

Pentachlorophenol 13 7.9* 

Trace Elements

Arsenic (trivalent) 69 36 

Arsenic (pentavalent) 2319* 

Copper 2.9 2.9 

Zinc 95 86 

* Insufficient data to develop criteria. Value presented is the lowest observed affect level.

NA: Not Available 

28 
RISK CHARACTERIZATION 

In its present condition and use, the Atlantic Wood site presents a continuing threat to aquatic habitats

based on examinations of potential pathways for contaminant migration, the toxicity hazard due to siterelated contaminants, and the estimated exposure levels. The south branch of the Elizabeth River, located 

immediately adjacent to the site, is of particular concern since it has received contamination through 

direct releases of creosote to the river. Since contamination in soil, groundwater, and sediment onsite 

represents a reservoir of source contamination, there is the continued potential for site-related 

contaminants to be transported off-site through surface water runoff, water discharge, and fugitive dust 

emissions from wind erosion. 

Since many of the contaminants of concern at Atlantic Wood will likely end up adsorbed to sediments, 

benthic and epibenthic organisms that have frequent contact with sediments will be most at risk. This 

includes benthic fish and invertebrates as well as those species that feed on benthic prey such as blue 

crab, weakfish, spot, and croaker. However, given the extreme levels of contamination, any organism that 

lives near the Atlantic Wood site may be at potential risk. 

Comparison Against Criteria and Standards

Based upon comparisons with EPA’s AWQC for the protection of aquatic organisms (Table 8), it is 

probable that aquatic species present in the inlet and along the shoreline of the site are at risk to acute and

chronic toxicity due to exposures of site-related contaminants. Maximum levels of ΣPAHs detected in 

groundwater exceed the AWQC concentrations by three orders of magnitude. Maximum levels of PCP

exceed AWQC by nearly two orders of magnitude. Given these extreme values, it is likely that AWQC

may be exceeded for these compounds even after groundwater are diluted upon discharge to surface 

waters2

. No sampling of surface waters has been conducted to verify this hypothesis. 

No criteria are presently available to evaluate the hazard of sediment contamination that are comparable

to the AWQC. For the purposes of this risk characterization, the data compiled by concentrations of 

contaminants of concern at the Atlantic Wood site that would be predictive of adverse biological impacts. 

Because no information was presented by Long and Morgan for PCP, the Apparent Effects Threshold

(AET) values (PTI, 1988a) have been presented for this compound. Only the maximum concentrations of

the contaminants of concern will be used. 

Based on the comparisons of the concentrations of the contaminants of concern in sediments with the

values compiled by Long and Morgan (1990), there is a significant potential for impact to aquatic species 

inhabiting on-site wetland areas, the inlet, and nearshore areas of the Elizabeth River adjacent to the site 

(Table 9). Impacts could occur at both the maximum estimated exposure level and the “more-likely”

(median) estimated exposure level. Because some of the data compiled by Long and Morgan (1990)

includes the results of chronic bioassays, it is probable that both chronic and acute effects could be

observed as a result of exposure of aquatic organisms to contaminated sediments from the Atlantic Wood

site.

 2 A modal of predicted exposure levels upon discharge of contaminated groundwater to the river was conducted and included as 

Appendix L in the RI. However, groundwater flow rates and contamination levels were not fully characterized; actual river flow 

data for calculation of dilution factors were not available; and, tidal flushing and contributory surface water sources were not 

accommodated. Since no ground-truthing of this model has been conducted, the model will not be incorporated in this 

assessment.

29 
Table 9. Comparison of the sediment concentrations (mg/kg) of the contaminants of concern to the data 

compiled by Long and Morgan (1990) and AETs (PTI, 1988a). 

Chemical Sediment Concentration Long & Morgan Data

Median Max ER-L ER-M OEAT AET 

Total PAH 1008 38,437 4 35 22 140 

PCP 15 67 ND ND ND 360 

Trace Elements

Arsenic N/D 374 33 85 50 57 

Copper N/D 1,350 70 390 300 390 

Zinc N/D 1,890 120 270 260 410 

Comparison of Estimated Exposure Levels with Toxicity Data

Sediment creosote toxicity data have been reported for laboratory and field studies, including field studies 

near the Atlantic Wood facility, using benthic macroinvertebrates and fish. In all the studies, creosote 

concentrations were expressed as ΣPAHs. The ΣPAH concentrations of these studies are compared to the 

maximum ΣPAH concentration in sediments of die inlet at the Atlantic Wood site (Table 10). The toxicity 

data cited in the literature reviewed is based upon the PAH composition of the sediments used in those 

studies. These data may differ from the PAH composition of sediments found at the Atlantic Wood site. 

However, for comparison purposes, it is assumed that the PAH compositions of the sediments are similar

enough to allow meaningful comparison. Based on the highest ΣPAH concentration of 38,437 mg/kg in

inlet sediments, the toxicity information indicates that there is a substantial risk to aquatic organisms in

the inlet and nearshore areas of the Elizabeth River adjacent to the Atlantic Wood site from exposure to 

site-related contaminants. Based upon the median value of 1,008 mg/kg, substantial risk to aquatic 

organisms is still indicated by the data in Table 10. 

The toxicity data show that a wide range of toxic responses could result from the levels of contamination 

at the site. Lethal toxicity could result from brief exposures to areas where extremely high concentrations 

of contaminants exist. Decreased species abundance and diversity may be expected, especially for benthic 

macroinvertebrate populations in direct contact with sediments, but also among the fish community.

Sublethal impacts, such as decreased reproductive success, immunosuppression, tissue damage, neoplasia,

and reduced scope for growth, could result from long-term exposure to contaminated sediments in this

stretch of the river. 

Some of these sublethal impacts, immunosuppression, for example, have been shown to be reversible

upon cessation of the toxic stress. The levels of PAH exposure at which there would be no observable 

adverse impacts have not been determined, however. 

30 
Table 10. Comparison of sediment PAH toxicity data from the literature with the maximum concentration 

of ΣPAH (mg/kg) in sediments from the inlet of 38,437 mg/kg. 

Species Total PAH Effect Reference

Invertebrates

Grass shrimp 5.8-65.5 Lethal and sublethal effects Alden & Butt, 1987 

Diporeia 100 Mortality Landrum et al., 1991 

Benthic invertebrates ≥177 Reduced colonization by polychaetes and

crustaceans; reduced biomass

Tagatz et al., 1983; Diaz, 

1985 

>844 Reduced colonization by mollusks Tagatz et al., 1983 

Fish

Spot 3 Induced MFO enzymes Van Veld et al., 1990 

Various species in the 

Elizabeth River

11.5 Reduced biomass Huggett et al., 1987 

Spot, weakfish, croaker 15.82 Cataracts Huggett et al., 1987 

Mummichog 61 Immunosuppression Faisal et al., in press

Hogchoker 100 Immunosuppression Weeks & Warinner, 1984 

Toadfish 154 Fin erosion Huggett et al., 1987 

Mummichog 2200 Presence of neoplasia Vogelbein et al., 1990 

Adverse Biological Effects

No specific studies were conducted at the Atlantic Wood site that relate on-site contaminant 

concentrations in sediments to adverse effects on aquatic organisms. However, a number of studies were 

conducted in the Elizabeth River next to the site that have related sediment concentrations of PAHs to

probable effects on aquatic organisms within the river. The results of these studies and the PAH levels at

which effects were associated have already been discussed in the Toxicity Assessment section and

summarized in Table 10. Only the types of effects observed will be reviewed here. 

Blue crabs collected in the south branch of the Elizabeth River just below and just above the Atlantic

Wood site were found to have elevated levels of PAHs in their muscle (up to 3,200 ppb) and 

hepatopancreas (up to 13,000 ppb) (Graves, 1989). The PAH composition in tissues of blue crab collected

from the majority of the sites in the Elizabeth River system was predominantly low molecular weight 

alkylated PAHs, such as alkyl naphthalenes. The pattern of PAH contamination in crab collected from the

Elizabeth River in the vicinity of the Atlantic Wood site changed to one more comparable with that of the

sediments in this area, showing an increase in unsubstituted PAHs. Spot from the area were found to have

induced MFO enzyme systems (Van Veld et al., 1990). These observations confirm the bioavailability, 

bioaccumulation, and bioresponse of resident organisms to existing PAH exposures adjacent to Atlantic

Wood. 

Macrophage activity in hogchoker from the vicinity were in a depressed state (Weeks and Warinner, 

1984). This depression of the cellular immune system could result in increased susceptibility to disease.

There are also reports of alterations in cancer defense mechanisms (Faisal et al., in press), indicating an

ongoing oncogenic process in fish from the area of the Atlantic Wood site. 

Tumors and other indications of tissue damage (fin erosion, cataracts, gill hyperplasia) have been

observed in fish collected near the site (Huggett et al., 1987). Grossly visible hepatic lesions occurred in 

93 percent of the mummichog collected next to Atlantic Wood (Vogelbein et al., 1990). 

Sediments from the Elizabeth River near the Atlantic Wood site were reported to rapidly produce lethal 

and sublethal effects in grass shrimp (Alden and Butt, 1987). 

Huggett et al. (1987) reported that the total number and biomass of several species of fish, including

hogchoker, toadfish, spot, weakfish, and croaker decreased from the mouth of the Elizabeth River to a 

station 2 km above the Atlantic Wood site. This decrease was associated with an increasing concentration 

31 
of PAHs in the sediments of the river. Additionally, higher incidences of fin erosion and cataracts in fish

were also associated with increasing sediment PAH concentration. 

These observations encompass a full range in levels of biological organization, from subcellular to

community parameters, and document existing adverse impacts due to PAH exposure. Based upon 

comparisons of the maximum and “more likely” exposure estimates with the reported concentrations 

associated with these impacts, sediments along the shoreline of the Atlantic Wood site may cause a wide

variety of acute and chronic effects on aquatic organisms. 

Joint Action of Chemical Mixtures

Potential receptors at the Atlantic Wood site are exposed to a diverse mixture of chemicals. The 

interactions among these chemicals may either increase or decrease the overall combined effects, or there 

may be no interaction at all. Another outcome of mixtures is that a compound that may not be particularly 

toxic by itself could become highly toxic in the presence of certain other chemicals. Predicting all the

possible interactions among the vast array of contaminants present onsite is not feasible; however, some 

general predictive indications from existing studies are available. 

As an approximation, chemicals that work by similar modes of toxic action typically will have a 

combined impact that is equal to the addition of all the individual impacts (an additive impact) when 

exposures are at high concentrations. For instance, exposure to two compounds each at LC10

concentration would give the same response as an LC20 concentration to either compound singularly. As 

exposure levels decrease, observations seem to indicate either the chemicals impart less joint action and

approach no interaction, or our ability to detect interactions are surpassed. Chemicals of differing toxic 

modes may have a less-than-additive effect, a more-than-additive effect, or may be antagonistic.

Predicting which outcome occurs for specific sets of chemicals based on chemical properties or structures

seems to have had limited success, and must be based at this point on empirical observation. 

There are two studies that supply empirical observations for contaminants of concern at the Atlantic 

Wood site. Landrum et al. (1991) studied the impact of PAH mixtures and Verma et al. (1981) has 

reported results with PCP. Landrum et al. (1991) exposed Diporeia (a burrowing amphipod) for up to 26

days to a mixture of PAHs with radio-labeled tracers at four molar concentrations and measured mortality 

and selected toxicokinetics. Their limited preliminary data suggests that the combined toxicity of a

mixture of individual PAH compounds is additive (i.e., the sum of toxicity for each of the individual

compounds). Verma et al. (1981) studied joint action of PCP, phenol, and dinitrophenol. Tests were made 

with all possible permutations of holding one or two chemical concentrations fixed at an LC0, while 

testing the LC50 of the other. All paired combinations were found to be more-than-additive (the combined

toxicity was greater than that predicted by adding the two individual toxicities). Most three-way

combinations were more-than-additive, one was additive, and two combinations were antagonistic. The 

two combinations that were antagonistic involved PCP held constant, while phenol or dinitrophenol were 

varied. 

Given the extreme levels of contamination observed in the inlet and shoreline at Atlantic Wood, 

exposures within these habitats may be considered as “acute” (equivalent or greater than those used in 

short-term lethality bioassays). They are also within the range used in the studies cited above. Data from 

these studies would then be clearly applicable to the exposure situations in these habitats. Results of these 

studies provide general indications that the overall toxicity of mixtures of PAHs, and mixtures of PCP 

with selected other contaminants, is the sum of the toxicities of all the individual compounds added 

together. Risk characterizations, which are determined for individual compounds, would therefore be 

considered only partial risks and underestimates of the total, overall risk of exposure to mixtures of 

contaminants.

32 
TERRESTRIAL ECOLOGICAL RISK ASSESSMENT 

POTENTIAL RECEPTORS 

The potential terrestrial receptors are those species that live within, or may transit through, the limited 

terrestrial habitat within the Atlantic Wood site. 

Terrestrial Habitats

The following represents a general characterization of the terrestrial habitat and fauna (mammalian, avian,

reptilian/amphibian) near the Atlantic Wood site. 

The site, located along the banks of the south branch of the Elizabeth River, is situated in an urbanized 

industrial complex. Overall habitat near the site is marginal to poor. Fences around the periphery of the 

site limit access for large mammals. 

Surrounding land use near the site is industrial and residential. Upland habitat and open space have been 

reduced to intermittent patches of deciduous forest and managed recreational or residential lands 

(playgrounds, golf courses, residential landscaping). Habitat providing a suitable home range (cover, 

food, water, migration corridors, and edge) for large mammals, particularly predators, is not available.

Numerous highways, roads, and railways below the site and north of Great Dismal Swamp form a 

physical barrier for mammals, less so for reptiles and amphibians, and are of no consequence to birds.

While not impassable, these features are an effective form of mortality on migratory mammal populations 

(Keel, personal communication, 1991). For this reason black bear, white tail deer, red fox, river otter,

mink, and bobcat are rarely observed outside Great Dismal Swamp near Portsmouth, Virginia. 

Surrounding upland habitat may still be utilized by songbirds (swallows, yellowthroats, etc.) for nesting 

and foraging. Also, it is probable that some waterfowl species may graze on available lawns and turf 

fields (Schwab, personal communication, 1991). 

Habitats of Special Concern

There are no significant or critical habitats located near the site. The closest critical habitats are two heron 

rookeries, great blue heron and yellow-crowned night heron, 5 km downstream near the Hampton Roads 

Army Terminal (USF&WS, 1980). 

Species

The site itself is largely devoid of natural vegetation, but the wetland areas on the site or in nearby off-site 

areas contain reed grass or saltbush communities, which are generally at least partially disturbed. 

No data were found describing the invertebrate species present at the site or in the wetlands. It would be

expected that normal soil and wetland invertebrate species would be supported in areas not occupied by

construction or paving, and not grossly contaminated. 

Available data on reptilian and amphibian species near the site are summarized in Table 11. Data were

available only from quite general sources. The species listed in Table 11 are presumed to be present at the

site based upon broad survey data. The many-lined salamander, Stereochilus marginatus, is a rare species 

that occurs in St. Julian Creek but has no state or federal legal status (Virginia Department of 

Conservation and Recreation, 1991). 

Mammals are limited near the site, but those present probably reside near the site for long durations,

increasing their exposure to contaminants. This may be particularly true for species that burrow into the

stream banks for shelter. The mammalian species near the site are typical of those species adapted to, and 

found near, urban environments. Muskrat, raccoon, marsh rabbit, and Norway rat are known to breed, 

raise young, and feed as adults near or on the site. Opossum, meadow vole, and marsh rice rat are also

33 
likely to breed near or on the site. It is suspected that feral cats are also present (Halbrook, personal 

communication, 1991). These populations are stable, probably because of limited available habitat rather 

than predation by carnivores. Recent studies indicated that muskrats taken from Paradise Creek were 

stressed (as indicated by elevated numbers of parasites, significantly higher concentrations of cadmium,

copper, and aluminum in their tissue, and lower body and liver weights), but did not identify conclusive

causes that could be linked with activities at the Atlantic Wood site (Halbrook, personal communication, 

1991). 

Survey data indicate that waterfowl, wading birds, shorebirds, raptors, and songbirds may be found 

utilizing the site during different parts of the year (Table 11). Waterfowl are considered the predominant

users of areas near the site. Nesting areas or rookeries near the site are extremely unlikely, especially with 

the abundance of superior habitat to the north and south. Utilization by avian species is primarily roosting 

and adult forage. 

Common waterfowl species are dabblers, black duck, mallard, and ruddy duck. These species feed

primarily on submerged and emergent aquatic vegetation near the shore of the river and in the adjacent

wetlands. Use by waterfowl is greatest during annual migrations (Beccasio, 1980). 

Wading birds most typically found are great blue heron, black-crowned night heron, yellow-crowned 

night heron, clapper rail, snowy egret, and great egret. It is possible that these species (except clapper rail) 

may roost periodically on piers, wharfs, and other waterfront structures. Great blue heron and yellowcrowned night heron are State-protected species. These latter two species typically feed on marine

invertebrates, reptiles, amphibians, and fish in shallow waters (Beccasio, 1980). 

34 
Table 11. Preliminary listing of terrestrial fauna utilizing habitat in the general vicinity of the Southern 

Branch of the Elizabeth River (Beccasio, 1980; USFWS, 1980; Niering, 1987; Keel, personal 

communication, 1991; Schwab, personal communication, 1991; Virginia Department of 

Conservation and Recreation, 1991). 

Common Name Scientific Name Breeding Juvenile Adult Forage

Reptiles/ Amphibians

Two-lined salamander Eurycea bislineata ♦

Mud turtle Kinosternon subrubrum ♦ ♦ ♦

Brown water snake Nerodia taxispilota ♦ ♦ ♦

Northern water snake Nerodia sipedon ♦ ♦ ♦

Mud salamander Pseudotriton montanus ♦

Pickerel frog Rana palustris ♦ ♦ ♦

Many-lined salamander Stereochilus marginatus ♦ ♦ ♦

Birds* 

Red-winged blackbird Agelaius phoeniceus ♦ ♦ ♦

Mallard duck Anas platyrhynchos ♦ ♦

Black duck Anas rubripes ♦ ♦

Great blue heron Ardes herodias ♦

Great egret Casmeroidius albus ♦

Marsh wren Cistothorus palustris ♦ ♦ ♦

Peregrine falcon Falco peregrinus ♦

Bald eagle Haliaetus leucocephalus ♦

Herring gull Larus argentatus ♦

Black-backed gull Larus marinus ♦

Snowy egret Leucophoyx thula ♦

Swamp sparrow Melospiza georgiana ♦ ♦ ♦

Black-crowned night heron Nycticorax nyticorax ♦

Yellow-crowned night heron Nycticorax violaceus ♦

Ruddy duck Oxyura jamaicensis ♦

Osprey Pandion haliaetus ♦

Double-crested cormorant Phalacrocorax auritus ♦

Clapper rail Rallus longirostris ♦

Mammals

Opossum Didelphis marsupialis ♦

Meadow vole Microtus pennsylvaicus ♦ ♦

Muskrat Odatra zibethica ♦ ♦ ♦

Marsh rice rat Oryzomys palustris ♦ ♦ ♦

Raccoon Procyon lotor ♦ ♦

Norway rat Rattus norvegicus ♦ ♦ ♦

Marsh rabbit Sylvilagus palustris ♦ ♦

Gray fox Urocyon cinereoargenteus ♦

* Habitat use by avian species is temporary, primarily for roosting and foraging, and variable depending on seasonal migrations.

It is probable that songbirds may be nesting near the site, but doubtful that any waterfowl, wading birds, shorebirds, or raptors

are also nesting.

35 
Three raptors, all listed as federally protected, periodically use habitat near the site. Peregrine falcon and 

osprey are the most common, and an occasional bald eagle has been observed. The utilization by these

species is thought to be infrequent (Schwab, personal communication 1991). 

Typical shorebirds include great black-backed gull, herring gull, and double-crested cormorant. Redwinged blackbird, swamp sparrow, and marsh wren are species most likely found near the site. 

EXPOSURE ASSESSMENT 

Contaminants of Concern

Identification of the contaminants of concern for the Atlantic Wood site is based on findings in the RI 

(KER, 1990). PAHs, PCP, PCDDs, PCDFs, arsenic, copper, and zinc have already been identified in the 

Aquatic Ecological Risk section. Screening approaches for contaminants of concern will not be discussed 

further in this section. 

Sources of Contaminants

The pattern of soil contamination is described in the Aquatic Ecological Risk section and is depicted in 

Figure 4. Sources will not be discussed further in this section. 

Fate and Transport Analysis

The environmental behavior of the contaminants of concern has already been discussed in the Aquatic 

Ecological Risk section and will not be discussed further here. 

Exposure Pathways

The primary sources of exposure to the contaminants of concern are through surface soils and inhalation. 

Ingestion and dermal absorption are pathways considered from contact with contaminated soils. 

Inhalation is an additional pathway of concern for burrowing animals. An exposure estimate was made for 

a representative mammal, bird, and amphibian through the oral exposure route. Two exposure scenarios 

were assumed: a realistic (low) exposure and an extreme exposure. 

Estimated Exposure Levels

In selecting representative target species, no information was found regarding the morphometrics and

dietary intake of species found near Atlantic Wood. Wildlife exposures were based on information for

species found in areas supporting forest vegetation (Table 12). The methodology used to determine the 

exposures is the same as that presented in the environmental impact statements prepared by the U.S. 

Forest Service (USFS) (1988). Representative species were selected based on their probable similarity in 

size and dietary intake to species that might be found at the Atlantic Wood site. 

Ingestion Doses 

Each representative species was assumed to feed on contaminated food items at the site. These dietary

amounts are listed in Table 12. In the realistic exposure scenario, each species is assumed to consume a 

portion of its daily intake in contaminated food items based on its body size. The percentage of food

contaminated (PFC) is based on the following formula: 

PFC - 100 x (1/BW)

0.2

, 

where BW is body weight (USFS, 1988). In the extreme case, each species’ entire daily intake is assumed

to consist of contaminated items. 

In addition to ingestion of contaminants through diet, birds and mammals can also ingest contaminants 

through preening of feathers and grooming of fur. The amount of contaminant ingested through this route 

was not calculated. 

36 
Dermal Exposure 

Although dermal exposures were not calculated, it should be noted that this is another route of

contaminant uptake. Dermal exposures are assumed to come primarily from contact with surface soils for 

mammals and birds, and from contact with surface soils and sediments for amphibians. Fur, feathers, and

scales afford varying degrees of protection against dermal exposure. It is likely that amphibians would 

receive a larger dose of a contaminant when exposed to the same concentration as a mammal or bird

because the moist, glandular skin of the amphibian serves to a large extent as a respiratory organ and is 

much more permeable than the skin of other animal classes (USFS, 1988). 

Inhalation Exposure 

At the Atlantic Wood site, contaminated airborne particulate matter resulting from wind erosion or 

construction practices would be the most likely scenario for inhalation exposure. As with any exposure

route, the dose received would be a function of concentration, frequency, and duration of exposure. With

inhalation exposure, particulate size is another factor that must be considered when estimating the dose an

animal receives. 

Contamination is known to be present in the surface and subsurface soils throughout the upland portions

of the site. In addition, soils and sediments in wetland areas of the site are also contaminated. Reptiles and 

mammals can be exposed to contamination from the site through the ingestion of food organisms that are 

contaminated and through the direct ingestion of contaminated soil on food and from cleaning activities. 

Birds can be exposed primarily through the ingestion of contaminated food from upland, wetland, and 

surface water habitats. Because of the extent of resource use, surface water-dependent birds (ducks and 

shorebirds) are considered to be most at risk. 

Four representative species were selected for the Atlantic Wood site: one omnivorous mammal, the cotton 

rat; two avian species, the kingfisher, a piscivore and the eastern bluebird, an omnivore; and one

insectivorous amphibian, the woodhouse toad. Because no morphometric or food consumption data were 

found for species that may actually be present at the site, the species listed above were selected based on 

their probable similarity to species found on site. The morphometric and food consumption data for the

selected species were used to calculate estimated daily doses from ingestion of contaminated food items. 

Two exposure scenarios were assumed, low and extreme exposures. In the low-dose scenario, the 

percentage of food contaminated was calculated using the percentage of food contaminated (PFC) 

equation, which is representative of a more realistic situation (USFS, 1988). In the extreme exposure, 100 

percent of the diet is assumed to be contaminated. In each case, the maximum concentration of

contaminant found in on-site soil was used as the exposure concentration. In the cases of the kingfisher 

and the woodhouse toad, the maximum concentration of ΣPAH found in inlet sediments was used in

calculating the estimated dose because these species are more likely to come in contact with sediments

than the other representative species. 

Estimated daily doses (EDD) of contaminants were calculated as follows: 

EDD = (Contaminant Concentration in mg/kg) (Daily food Intake is g/dy) 

(Body Weight in g) 

The estimated daily doses are summarized in Table 12. 

Uncertainty Analysis

There is a level of uncertainty associated with any approach employed to determine potential risk to

potential receptors posed by exposure to contaminants. This uncertainty usually stems from lack of

knowledge of the true exposure regimes, lack of information on the health risk posed by a contaminant to

the specific receptors of concern, or lack of detailed information on the biology of potential receptors. The 

overall, cumulative risk posed by exposure to multiple contaminants, either simultaneously or 

37 
sequentially, also carries a great deal of uncertainty with it due to the lack of data on multiple exposure 

toxicity. In this risk assessment, the uncertainty of estimating exposure levels is due to a lack of sitespecific data regarding the use of the Atlantic Wood site by terrestrial species. The concentrations used in

calculating the estimated daily doses are those found in soil and sediments rather than in the actual food 

items selected by the representative species. In foraging, it is likely that animals will ingest a certain

percentage of contaminated soil, but this probably represents only an incidental amount in most cases. 

The PFC equation only takes into account the body weight of the animal for estimating the percentage of

food that is contaminated and does not account other factors such as residence tune within a particular

area or the proportion of the daily food intake that might be consumed within a given location. This factor 

leads to potential biases in determining exposure estimates. For instance, it is recognized that one would 

expect birds to have a larger home range and thus consume proportionately less food from a given area 

than a small mammal, and thus birds have smaller exposure levels. This subjective assumption is not 

reflected in the estimated exposure levels calculated by the PFC equation (Table 12). Because the 

equation is quite general and relies only on one parameter, it is limited in its application. However,

because so little is known about the specific species that use the Atlantic Wood site and the behaviors of

those species, it is necessary to base estimates on the data that are available. Because the PFC equation is 

objective, and does not arbitrarily specify the percentage of food consumed on site by each target species, 

this method chosen over more subjective approaches to estimating exposure. 

The estimated daily doses for the low-dose scenario in the cotton rat and the woodhouse toad, though 

conservative, probably represent a more realistic situation. Small species of mammals and amphibians 

tend to have smaller home ranges and will spend more time in a given area than larger species (USFS,

1988). For this reason, small mammals and amphibians at the Atlantic Wood site are more likely to spend 

substantial portions of their lives on the site and consume a greater percentage of their diet there. 

From the limited information available, it does not appear that the site is used extensively by avian

species. Though the site is located along the corridor of the Atlantic Flyway, it is doubtful the site is a 

preferred habitat for bird species. Some incidental exposure could occur, but the estimated daily dose for 

the low exposure scenario is most likely excessively conservative and overestimates the actual exposure 

that avian species would encounter at the site. 

38 
Table 12. Representative wildlife, their associated biological parameters, and the estimated dose (mg/kg 

bw/day) of contaminant received through ingestion (does not include dose received from 

grooming and preening activities) (USFS, 1988). 

Percentage of food 

contaminated 

Species Chemical1 Body weight 

(g) 

Daily intake

(g) 

Low Extreme 

Low 

Dose2

Extreme 

Dose2

MAMMALIAN 

Cotton rat 

(omnivorous) 156 45 36 100 

Total PAH 1,376 3,322 

PCP 100 280 

TCDD 0.001 0.004 

Arsenic 51 143 

Copper 1,015 2.321 

Zinc 2,118 5,885 

AVIAN

Kingfisher 

(piscivorous) 250 60 33 100 

Total PAH 3,044 9,225 

PCP 77 233 

TCDD 0.001 0.003 

Arsenic 39 119 

Eastern bluebird

(omnivorous) 29 6 51 100 

Total PAH 1,398 2.741 

PCP 100 200 

TCOO 0.001 0.003 

Arsenic 52 102 

AMPHIBIAN 

Woodhouse toad

(insectivorous) 22 5 54 100 

Total PAH 4,717 8,736 

1 The concentration of chemical used in calculating dose was the highest surface soil concentration found on site. For the 

kingfisher and the woodhouse toad, the highest ΣPAH concentration found in on-site sediments was used in calculating the 

estimated daily dose because these species are more likely to come in contact with sediments than the other representative

species. 

2 Estimated daily doses were calculated only for those chemicals for which toxicity information was available for the species

listed.

TOXICITY ASSESSMENT 

This section summarizes the toxicity of the contaminants of concern at the Atlantic Wood site to

terrestrial species. For the purposes of this study, terrestrial species will include mammals, birds,

amphibians, and reptiles. 

Reported Effects of Contaminants of Concern

The terrestrial toxicity assessment summarizes the findings of laboratory studies that indicated the 

toxicity to wildlife of the contaminants of concern found in surface soils and sediments at the Atlantic 

Wood site. The term wildlife, as used here, refers to mammals, birds, reptiles, and amphibians. In most 

cases, laboratory studies of domestic animals have been used because of a lack of studies specifically on 

wildlife. The results of domestic animal studies are considered to be representative of the effects that

would occur in similar species in the wild. Toxicity information for each of the contaminants of concern 

39 
has been tabularized (Tables 13 through 18) to include data for mammalian, avian, reptilian, and 

amphibian species. The toxicity information has been subdivided into acute and chronic studies. 3

 

Table 13. Summary of reported levels for PAH toxicity in wildlife (EPA, 1986d; Eisler, 1987). 

Animal Class Exposure Dose Effects

MAMMALIAN 

Acute 50-2,000 mg/kg bw LD-50 

Chronic 80-100 mg/kg bw Reproductive impairment 

4x10-5 - 3,300 mg/kg bw Carcinogenesis

20-250 mg/kg diet Not reported Carcinogenesis

AVIAN

Acute 1,261 mg/kg bw LC-50 

Chronic 4,000 mg/kg bw No effect 

0.002 μg/egg Teratogenic effects 

REPTILIAN &

AMPHIBIAN 

Acute 40 mg/kg bw Induction of mixed-function 

oxidase activity

Chronic 1.5 mg (implanted into 

abdominal cavity)

Production of lymphosarcoma

Table 14. Summary of reported levels for PCP toxicity in wildlife (EPA, 1986c; Eisler, 1989)

Animal Class Exposure Dose Effects 

MAMMALIAN 

Acute

Laboratory rodents 27-300 mg/kg bw LD-50 

Eastern chipmunk 138 mg/kg bw LD-50 

Chronic

Laboratory rodents 20-500 mg/kg feed 1.2-30 mg/kg bw/day Disruption of liver enzyme

activity and reduced weight

gain after 8-month exposure 

200 mg/kg diet 13 mg/kg bw/day Teratogenic 

AVIAN

Acute

Mallard & ring-necked

pheasant 

380-504 mg/kg bw LD-50 

Japanese quail 5,139 mg/kg diet Not reported LC-50 

Chronic 

Domestic chicken 1-1,000 mg/kg diet Not reported Liver abnormalities 

AMPHIBIAN & 

REPTILIAN 

No studies were found 

 3 Designations of “acute” and “chronic” doses employed are reproduced as stated in the original references.

40 
Table 15. Summary of reported levels for 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) toxicity in

wildlife (Eisler, 1986; EPA, 1986e). 

Animal Class Exposure Dose Effects

MAMMALIAN 

Acute

Laboratory rodents 0.6-284 μg/kg LD-50 

0.01 -3.57 μg/kg/day Decreased body weight toxic 

hepatitis & porphyria after 4 to 

13 weeks exposure 

0.01-3 μg/kg/day Teratogenic effects after

maternal exposure during days

0.1 μg/kg/day Carcinogenesis

AVIAN

Acute

Quail, dove, mallard & 

chicken 

l5-810 μg/kg/day LD-50 

Chronic

Chicken 1 -10 μg/kg/day of mixed Death 

PCDDs Liver enlargement & necrosis 

Pericardial, subcutaneous & 

peritoneal edema 

REPTILIAN &

AMPHIBIAN 

No studies were found 

Table 16. Summary of reported levels for arsenic (inorganic) toxicity to wildlife (Eisler, 1988). 

Animal Class Exposure Dose Effects

MAMMALIAN 

Acute

Various mammalian species 35-100 mg/kg bw 

(calcium arsenite) 

=13-38 mg As/kg bw

LD-50 

1-25 mg/kg bw 

(sodium arsenite) 

=0.6-14.5 mg As/kg bw 

Lethal 

3-250 mg/kg bw 

(As trioxide) 

=2.3-190 mg As/kg bw 

Lethal 

Chronic

Mouse 5 mg/kg diet for 3 generations

(sodium arsenite) = 2.9 mg 

As/kg diet 

Not reported Reduced litter size 

AVIAN

Acute

Mallard, quail & pheasant 47.6-386 mg/kg bw (sodium 

arsenite) =28-224 mg As/kg 

bw

LD-50 

Mallard 1,000 mg/kg diet (sodium 

arsenite) =580 mg As/kg diet 

Not reported LD-50 

Chronic No studies were found 

REPTILIAN &

AMPHIBIAN 

No studies were found 

41 
Table 17. Summary of reported levels for copper toxicity to wildlife (EPA, 1986b). 

Animal Class Exposure Dose Effects 

MAMMALIAN 

Acute No studies were found 

Chronic

Laboratory rodents 5,000 mg/kg diet 

(Cu acetate) 

= 1,600 mg Cu/kg diet 

Not reported Accumulation in liver &

kidney

3,000-4,000 mg/kg diet 

(Cu sulfate) 

=1,195-1,593 mg Cu/kg diet

155-207 mg/kg bw/day Embryonic malformations 

AVIAN No studies were found 

REPTILIAN &

AMPHIBIAN 

No studies were found 

Table 18. Summary of reported levels for zinc toxicity to wildlife (EPA, 1986f)A

Animal Class Exposure Dose Effects

MAMMALIAN 

Acute No studies were found 

Chronic

Laboratory rodents 5,000 ppm Zn in drinking

water 

850 mg/kg bw/day No observed effect 

30,000 ppm in diet (Zn sulfate) 

=12,150 ppm Zn

Approx. 950-1,880 mg/kg 

bw/day

Decreased food intake 

Retarded growth

Hematologic abnormalities 

5,000-10,000 ppm Zn in diet 250-500 mg/kg bw/day Mortality 

Reproductive impairment 

Reduced growth

AVIAN No studies were found 

REPTILIAN &

AMPHIBIAN 

No studies were found 

Site-Specific Toxicity Tests

No site-specific bioassays have been conducted with terrestrial species. 

Existing Toxicity-Based Criteria and Standards

There are currently no existing toxicity-based criteria or standards for soils. 

RISK CHARACTERIZATION 

As with the aquatic species, insufficient data are available regarding the presence of terrestrial species

near the site to allow a quantitative risk assessment to be performed. In addition, no toxicity data are 

available for species that have been identified near the site. Because of these limitations, the risk

assessment will be limited to a more general consideration of potential toxicity to mammals (primarily

rodents), birds, reptiles, and amphibians. 

In selecting representative target species, no information was found regarding the morphometrics and

dietary intake of species found in the vicinity of Atlantic Wood. Species selection was based on a list of

species presented in an environmental impact statement prepared by the USFS (1988) which listed 

42 
mammalian, avian, reptilian, and amphibian species; their estimated body weights; and their estimated

daily dietary intake. Representative species were selected based on their probable similarity in habitat, 

size, and dietary habits to species that might be found at the Atlantic Wood site. 

Comparison of Estimated Exposure Levels with Toxicity Data

In this section, the estimated daily doses of the chemicals of concern received by representative target 

species through ingestion are compared to laboratory toxicity data (ingestion only) (Table 19). Both

chronic and acute laboratory toxicity data were used for comparison to the estimated daily doses. Daily

doses due to inhalation and dermal exposures were not calculated because there is insufficient data to 

make such calculations and because the estimated daily doses calculated due to ingestion most likely

overestimate the doses being received by receptor species using the site. 

Based on the low-dose scenario of the estimated daily doses calculated for the chemicals of concern at the

Atlantic Wood site, mammalian, avian, and amphibian species onsite are at risk due to exposure to these 

chemicals. Mammalian species could suffer chrome and acute toxicity from ingestion of PAHs, PCP, 

PCDDs and PCDFs, and arsenic. Concentrations of copper and zinc are high enough so that mammals

could suffer chronic toxicity from long-term exposure to these substances. 

Avian species are potentially at risk from both chronic and acute toxicity from exposures to the PAHs, 

PCDDs and PCDFs, and arsenic. No toxicity data were available describing the toxicity of copper and

zinc to avian species, so the risk of these substances to avian species cannot be addressed. 

Amphibian species could receive both chronic and acute doses of the PAHs. No toxicity data were found 

describing the toxicity of the other chemicals of concern to either amphibians or reptiles, so the risk to 

these species associated with exposure to the other chemicals will not be addressed. 

Adverse Biological Effects

Very limited information was available describing habitats on the site. A wetland delineation study 

conducted by KER, Inc. (1990) reported five wetland areas on the site described as being in a disturbed 

condition and of very low ecological importance. No information was available describing presence or 

absence of key species, population assessments of key species, or community indices. No studies have

been conducted relating contaminant concentrations at the site with adverse effects in terrestrial species. 

Joint Action of Chemical Mixtures

Joint action of chemicals mixtures has already been discussed in the Aquatic Ecology Risk section. Such

action will not be discussed further in this section.

43 
Table 19. Comparison of laboratory toxicity data with the estimated daily doses that the selected target

species could receive at the Atlantic Wood site through exposure to site soils or sediments 

(Eisler 1986, 1987, 1988, 1989; EPA 1986 b,c,d,e,f). All concentrations are expressed as 

mg/kg body weight unless otherwise noted. 

Representative Species & Laboratory Toxicity Data Estimated Daily Doses

Chemical Acute Chronic Low Dose Extreme Dose

MAMMALIAN 

Cotton Rat

Total PAH 50-2,000 4x10-5-3,300 1,376 3,822 

PCP 27-300 1.2-30 100 280 

TCDD 0.0006-0.284 (1-357)x10-5 0.001 0.004 

Arsenic 0.6-190 2.9 mg/kg diet 51 143 

Copper No data 155-207 1,015 2,821 

Zinc No data 250-1,880 2,118 5,885 

AVIAN

Kingfisher

Total PAH* 1,261 4,000 3,044 9,225 

PCP 380-504 1-1,000 mg/kg diet 77 233 

TCDD 0.015-0.81 0.001-0.01 0.001 0.003 

Arsenic 48-386 39 119 

Eastern Bluebird

Total PAH 1,261 4,000 1,398 2,741 

PCP 380-504 1-1,000 mg/kg diet 100 200 

TCDD 0.015-0.81 0.001-0.01 0.001 0.003 

Arsenic 48-386 No data 52 102 

AMPHIBIAN 

WoodhouseToad

Total PAH* No data 40 4,717 8,736 

* For these species, the highest sediment concentration of ΣPAHs was used in calculating the estimated daily dose.

REFERENCES 

Alden, R.W., and A.J. Butt. 1987. Statistical classification of the toxicity and polynuclear aromatic 

hydrocarbon contamination of sediments from a highly industrialized seaport. Environmental Toxicology 

and Chemistry 6: 673-684.

Beccasio, A.D., G.H. Weissberg, A.E. Redfield, R.L. Frew, W.M. Levitan, J.E. Smith, and R.E. Godwin. 

1980. Atlantic coast ecological inventory user's guide and information base. Washington, D.C: U.S. Fish 

and Wildlife Service, Biological Services Program. 

Choudhury, H., J. Coleman, C.T. Rosa, and J.F. Stara. 1986. Pentachlorophenol: health and 

environmental effects profile. Toxicology and Industrial Health 2: 483-574. 

Chemical Information Systems, Inc. (CIS). 1986. Oil and Hazardous Materials/Technical Assistance Data 

System. 

Clement Associates. 1985. Chemical, Physical, and Biological Properties of Compounds present at 

Hazardous Waste Sites. Washington, D.C.: U.S. Environmental Protection Agency. 

Diaz, R.J. 1985. Macrobenthos of the Elizabeth River in relation to hydrocarbon contamination of the 

sediments. Final Report to the Coastal Energy Impact Program. Virginia Institute of Marine Science, 

44 
College of William and Mary. Gloucester Point, VA. 

Eisler, R. 1986. Dioxin hazards to fish, wildlife, and invertebrates: a synoptic review. U.S. Fish and 

Wildlife Service Biological Report 85 (1.8). 37 pp. 

Eisler, R. 1987. Polycyclic Aromatic Hydrocarbon Hazards to Fish, Wildlife, and Invertebrates: A 

Synoptic Review. U.S. Fish and Wildlife Service Biological Report 85 (1.11). 81 pp. 

Eisler, R. 1988. Arsenic hazards to fish, wildlife, and invertebrates: a synoptic review. U.S. Fish and 

Wildlife Service Biological Report 85 (1.12). 92 pp. 

Eisler, R. 1989. Pentachlorophenol Hazards to Fish, Wildlife, and Invertebrates: A Synoptic Review. U.S. 

Fish and Wildlife Service Biological Report 85 (1.17). 72 pp. 

ESC. 1988. Remedial Investigation/Feasibility Study Work Plan and Appendices for the Atlantic Wood 

Industries, Inc. Site. Vienna, VA: Environmental Strategies Corporation. 

Faisal, M., B.A. Weeks, W.K. Vogelbein, and R.J. Huggett. In press. Evidence of aberration of the natural 

cytotoxic cell activity in Fundulus heteroclitus (Pisces: Cyprinodontidae) from the Elizabeth River, 

Virginia. Virginia Institute of Marine Science, School of Marine Science, College of William and Mary, 

Gloucester Point, VA. 

Firestone, D., J. Ress, N.C. Brown, R.P. Barron, and J.N. Daraico. 1972. Determination of 

polychlorinated dibenzo-p-dioxins and related compounds in commercial chlorophenols. Journal 

Association of Official Analytical Chemists 5: 85-92.

Gillingham, Lewis. Virginia marine resources commission, Newport News, Virginia, personal 

communication, October 5,1989. 

Graves, J. 1989. Elizabeth River Long-Terra Monitoring Program - Phase 1,1989: Analysis of organic 

pollutants in sediments and blue crab (Callinectes sapidus) tissues. Final Report to the Virginia State 

Water Control Board, Richmond, VA. 

Halbrook, Richard S. 1991. Wildlife biologist, Oak Ridge National Laboratory, Oak Ridge, TN. Personal 

communication, March 5,1991. 

Hargis, W.J., M.H. Roberts, and D.E. Zwerner. 1984. Effects of contaminated sediments and sedimentexposed effluent water on estuarine fish: acute toxicity. Marine Environmental Research 14: 337-354. 

Huggett, R.J., M.E. Bender, and M.A. Unger. 1987. Polynuclear aromatic hydrocarbons in the Elizabeth 

River, Virginia. In: Dickson, K.L., A.W. Maid, and W.A. Brungs (eds.). Fate and Effects of SedimentBound Chemicals in Aquatic Systems. Proceedings of the Sixth Pellston Workshop, Florissant, CO, 

August 12-17, 1984. Soc. Environ. Toxicol Chem., Spec. Pub. Ser. pp 327-341. 

Jensen, S. and L. Renberg. 1973. Chlorinated dimers present in several technical chlorophenols used as 

fungicides. Environmental Health Perspect 5: 37-39. 

Johnson, R.L., P.J. Gehring, R.J. Kociba and B.A. Schwetz. 1973. Chlorinated dibenzodioxin and 

pentachlorophenol. Environmental Health Perspect 5: 171-175. 

Johnson, L.L., E. Casillas, T.K. Collier, B.B. McCain, and U. Varanasi. 1988. Contaminant effects on 

ovarian maturation in English sole (Parophrys vetulus) from Puget Sound, Washington. Canadian Journal 

of Fisheries and Aquatic Sciences 45: 2133-2146. 

Keel, Ralph. 1991. Staff biologist, Great Dismal Swamp National Wildlife Refuge, Suffolk, VA. Personal 

communication, February 28, 1991. 

Keystone Environmental Resources, Inc. (KER). 1990. Remedial Investigation Report of Atlantic Wood 

Industries, Inc., Portsmouth, Virginia Site. Prepared for Atlantic Wood Industries, Inc., Savannah, GA. 

45 
Konasewich, D.E., P.M. Chapman, E. Gerencher, G. Vigers, N. Treloar. 1982. Effects, pathways, 

processes, and transformation of Puget Sound contaminants of concern. NOAA Technical Memorandum

OMPA-20. Boulder, CO. National Oceanic and Atmospheric Administration. 357 pp. 

Kuehl, D.W., P.M. Cook, and A.R. Batterman. 1986. Uptake and depuration studies of PCDDs and 

PCDFs in freshwater fish. Chemosphere 15: 2023-2026. 

Kuehl, D.W., P.M. Cook, A.R. Batterman, D. Lothenbach and B.C. Butterworth. 1987. Bioavailability of 

polychlorinated dibenzo-p-dioxins and dibenzofurans from contaminated Wisconsin River sediment to 

carp. Chemosphere 16: 667-679. 

Landrum, P.P., B.J. Eadie and W.R. Faust. 1991. Toxicokinetics and toxicity of a mixture of sedimentassociated polycyclic aromatic hydrocarbons to the amphipod Diporeia sp. Environmental Toxicology

and Chemistry 10: 35-46. 

Long, E.R. and M.F. Buchman. 1989. An evaluation of candidate measures of biological effects for the 

National Status and Trends Program. NOAA Technical Memorandum NOS OMA 45. Seattle, WA. 

National Oceanic and Atmospheric Administration. 105 pp. 

Long, E.R. and L.G. Morgan. 1990. The Potential for Biological Effects of Sediment-sorbed 

Contaminants Tested in the National Status and Trends Program. NOAA Technical Memorandum NOS 

OMA 52. Seattle, WA. National Oceanic and Atmospheric Administration. 175 pp. + appendices. 

Malins, D.C., M.M. Krahn, M.S. Myers, L.D. Rhodes, D.W. Brown, C.A. Krone, B.B. McCain, and S.L. 

Chan. 1985. Toxic chemicals in sediments and biota from a creosote-polluted harbor: relationships with 

hepatic neoplasms and other hepatic lesions in English sole. Carcinogenesis 10: 1463-1469. 

Malins, D.C., B.B. McCain, M.S. Myers, D.W. Brown, M.M. Krahn, W.T. Roubal, M.H Schiewe, J.T. 

Landahl, and S.L. Chan. 1987. Field and laboratory studies of the etiology of liver neoplasms in marine 

fish from Puget Sound. Environmental Health Perspectives 71: 5-16. 

McGreer, E.R. 1979. Sublethal effects of heavy metal contaminated sediments on the bivalve Macoma

balthica (L.). Marine Pollution Bulletin 10(9): 259-262.

Merrill, E.G. and T.L. Wade. 1985. Carbonized coal products as a source of aroma hydrocarbons to 

sediments from a highly industrialized estuary. Environmental Science and Technology 19: 597-603. 

Mix, M.C. 1979. Chemical carcinogens in bivalve mollusks from Oregon estuaries. EPA-600/3-79-034. 

Gulf Breeze, FL: U.S. Environmental Protection Agency. 33 pp. 

Mothershead, R.F. II, R.C. Hale. Influence of Ecdysis on the accumulation of polycyclic aromatic 

hydrocarbons in field exposed blue crabs (Callinectes sapidus). Gloucester Point, VA. Virginia Institute 

of Marine Science, College of William and Mary. 24 pp. 

Moyle, P.B. and J.J. Cech, Jr. 1982. Fishes: An introduction to Ichthyology. Prentice-Hall, Inc. New 

Jersey. 

Mueller, J.G., P.J. Chapman, and P.H. Pritchard. 1989. Creosote-contaminated sites: their potential for 

bioremediation. Environmental Science and Technology 23: 1197-1201. 

Niering, William A. 1987. Wetlands. The Audubon Society Nature Guides. Alfred A. Knopf, Inc., New 

York. 

Niimi, A.J. and C.Y. Choy. 1983. Laboratory and field analysis of pentachlorophenol (PCP) accumulation

by salmonids. Water Research 17: 1791-1795. 

Phelps, H.L., J.T. Hardy, W.H. Pearson, and C.W. Apts. 1983. Clam burrowing behaviour: inhibition by 

Cu-enriched sediment. Marine Pollution Bulletin 14(12): 452-455. 

PTI. 1988a. Preliminary Literature Review of the Aquatic Transport, Fate, and Effects of Creosote; and 

46 
Recommendations for Chemical and Biological Studies. Tech. Memo. 88-4. Seattle, Washington: 

National Oceanic and Atmospheric Administration. 

PTI. 1988b. The briefing report to the Environmental Protection Agency Science Advisory Board: The 

Apparent Effects Threshold approach. Washington, D.C.: Environmental Protection Agency. 

Qasim, S.R., A.T. Armstrong, J. Corn, and B.L. Fordan. 1980. Quality of water and bottom sediments in 

the Trinity River. Water Resources Bulletin 16(3): 522-531. 

Roberts, M.H., Jr., W.J. Hargis, Jr., C.J. Strobel, and P.P. DeUsle. 1989. Acute toxicity of PAH 

contaminated sediments to the estuarine fish, Leiostomus xanthurus. Bulletin of Environmental

Contamination and Toxicology 42:142-149. 

Rubinstein, N.I., RJ. Pruell, B.K. Taplin, J.A. LiVolsi, and C.B. Norwood. 1990. Bioavailability of 

2,3,7,8-TCDD, 2,3,7,8-TCDF and PCBs to marine benthos from Passaic River sediments. Chemosphere

20: 1097-1102. 

Schwab, Donald. 1991. Field biologist, Virginia Department of Game and Inland Fisheries, Wildlife 

Division. Suffolk, VA. Personal communication, February 27, 1991.

Swartz, R.C., P.P. Kemp, D.W. Schults, and J.O. Lamberson. 1988. Effects of mixtures of sediment 

contaminants on the marine infaunal amphipod, Rhepoxynius abronius. Environmental Toxicology and

Chemistry 7: 1013-1020. 

Tagatz, M.E., J.M. Ivey, and M. Tobia. 1978. Effects of Dowicide G-ST on development of experimental 

estuarine macrobenthic communities. In: K.R. Rao, (ed.). Pentachlorophenol: chemistry, pharmacology, 

and environmental toxicology. pp 157-163. 

Spitzbergen et al., 1991 (Page 27). 

Swartz, R.C., P.F. Kemp, D.W. Schults, and J.O. Lamberson. 1988. Effects of mixtures of sediment 

contaminants on the marine infaunal amphipod, Rhepoxynius abronius. Environmental Toxicology and

Chemistry 7: 1013-1020. 

Tagatz, M.E., J.M. Ivey, and M. Tobia. 1978. Effects of Dowicide G-ST on development of experimental 

estuarine macrobenthic communities. In: K.R. Rao, (ed.). Pentachlorophenol: chemistry, pharmacology, 

and environmental toxicology. pp 157-163. 

Tagatz, M.E., G.R. Plaia, C.H. Deans, and E.M. Lores. 1983. Toxicity of creosote-contaminated sediment 

to field- and laboratory-colonized estuarine benthic communities. Environmental Toxicology and 

Chemistry 2: 441-450. 

Travelstead, J. 1989. Associate Commissioner for Fisheries Management, Virginia Marine Resources 

Commission, Personal communication, October 18, 1989. 

U.S. EPA. 1980. Ambient water quality criteria for pentachlorophenol. Washington, D.C.: U.S. 

Environmental Protection Agency, Office of Water Regulations and Standards, Criteria and Standards 

Division. EPA 440/5-80-065. 

U.S. EPA. 1985. Ambient water quality criteria for Cu - 1984. Washington, DC: U.S. Environmental 

Protection Agency, Office of Water Regulations and Standards, Criteria and Standards Division. EPA 

440/5-84-031.

U.S. EPA. 1986b. Health effects assessment for Cu. EPA/540/1-86-025. Cincinnati, OH: U.S. 

Environmental Protection Agency, Office of Research and Development, Office of Health and 

Environmental Assessment, Environmental Criteria and Assessment Office. 

U.S. EPA. 1986c. Health effects assessment for pentachlorophenol. EPA/540/1-86-043. Cincinnati, OH: 

U.S. Environmental Protection Agency, Office of Research and Development, Office of Health and 

47 
Environmental Assessment, Environmental Criteria and Assessment Office. 

U.S. EPA. 1986d. Health effects assessment for polycyclic aromatic hydrocarbons (PAHs). EPA/540/1-

86-013. Cincinnati, OH: U.S. Environmental Protection Agency, Office of Research and Development, 

Office of Health and Environmental Assessment, Environmental Criteria and Assessment Office. 

U.S. EPA. 1986e. Health effects assessment for 2,3,7,8-TCDD. EPA/540/1-86/044. Cincinnati, OH: U.S. 

Environmental Protection Agency, Office of Research and Development, Office of Health and 

Environmental Assessment, Environmental Criteria and Assessment Office. 

U.S. EPA. 1986f. Health effects assessment for Zn (and compounds). EPA/540/1-86-048. Cincinnati, OH: 

U.S. Environmental Protection Agency, Office of Research and Development, Office of Health and 

Environmental Assessment, Environmental Criteria and Assessment Office. 

U.S. EPA. 1986g. Quality Criteria for Water. 440/5-86-001. Washington, D.C.: U.S. Environmental 

Protection Agency, Office of Water Regulations and Standards, Criteria and Standards Division. 

U.S. EPA. 1990. Background document to the integrated risk assessment for dioxins and furans from

chlorine bleaching in pulp and paper mills. Washington, D.C.: U.S. Environmental Protection Agency, 

Office of Toxic Substances. EPA 560/5-90-014. 

U.S. Fish and Wildlife Service. 1980. Atlantic Coast Ecological Inventory Map. Norfolk, VA. 36076-A1-

E1-250. 

U.S. Fish and Wildlife Service. 1989. 

U.S. Forest Service. 1988. Final Environmental Impact Statement: Vegetation Management in the Coastal 

Plain/Piedmont, Volume II. Appendix A: Risk assessment for the use of herbicides in the southern region, 

United States Department of Agriculture, Forest Service. 

Van Veld, P.A., D.J. Westbrook, B.R. Woodin, R.C. Hale, C.L. Smith, R.J. Huggett and J.L. Stegeman. 

1990. Induced cytochrome P-450 in intestine and liver of spot (Leiostomus xanthurus) from a polycyclic 

aromatic hydrocarbon contaminated environment. Aquatic Toxicology 17: 119-132. 

Varanasi, U., W.L. Reichert, J.E. Stein, D.W. Brown, and H.R. Sanborn. 1985. Bioavailability and 

biotransformation of aromatic hydrocarbons in benthic organisms exposed to sediment from an urban 

estuary. Environmental Science and Technology 19: 836-841. 

Varanasi, U., J.E. Stein, and M. Nishiraoto. 1989. Biotransformation and disposition of polycyclic 

aromatic hydrocarbons (PAH) in fish. In: U. Varanasi (ed.), Metabolism of polycyclic aromatic 

hydrocarbons in the aquatic environment, p. 93-149. CRC Press, Inc., Boca Raton, FL. 

Verma, S., S. Rani, and R. Dalela. 1981. Synergism, antagonism, and additivity of phenol, 

pentachlorophenol, and dinitrophenol to a fish (Notopterus notopterus). Archives of Environmental 

Contamination and Toxicology 10: 365-370. 

Virginia Department of Conservation and Recreation, 1991. Letter communication with Kennedy H 

Clark, Environmental Review Coordinator. March 6,1991. Richmond, VA. 

Virginia Institute of Marine Sciences (VIMS). 1989. Trawl survey database for juvenile fishes: 1975-

1985, Gloucester Point, VA. 

Vogelbein, W.K., J.W. Fournie, P.A. Van Veld, and R.J. Huggett. 1990. Hepatic neoplasms in the 

mummichog Fundulus heteroclitus from a creosote-contaminated site. Cancer Research 50: 5978-5986. 

Walker et al., 1991 (Page 27) 

Weeks, B.A. and J.E. Warinner. 1984. Effects of toxic chemicals on macrophage phagocytosis in two 

estuarine fishes. Marine Environmental Research 14: 327-335. 

48 
Weeks, B.A. and J.E. Warinner. 1986. Functional evaluation of macrophages in fish from a polluted 

estuary. Vetinary Immunology and Immunopathology 12: 313-320. 

Weeks, B.A., J.E. Warinner, P.L. Mason, and D.S. McGinnis. 1986. Influence of toxic chemicals on the 

chemotactic response of fish macrophages. Journal of Fish Biology 28: 653-658.

Wright, M., Sanitarian Manager. Virginia Department of Shellfish Sanitation, Richmond, VA, Personal 

communication, January 10, 1991. 

Zieske, G. 1991. Atlantic Wood Industries, Inc. Portsmouth, Virginia NPDES Biomonitoring Program, 

#VA0004189, Outfalls #001, #002, and #002, 48-hour static acute definitive toxicity bioassays using 

Mysidopsis bahia and Cyprinodon variegatus August 21-23m 1991. Presented Savannah Laboratories and 

Environmental Services, Inc. Gainesville, FL. Alvarez, Lehman and Associates, Inc. 45 pp. 

49