ACKNOWLEDGEMENTS

ii

Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

DECEMBER

Toxic Contaminants in the Chesapeake Bay and 

its Watershed: Extent and Severity of Occurrence 

and Potential Biological Effects

TECHNICAL REPORT 2012

U.S. Geological Survey

U.S. Fish and Wildlife Service

Chesapeake Bay Field Office

Annapolis, Maryland

H S E L L F I S H

•

F I

S

H

•

S

E

D I M

E

N

T

•

W

I

L D L I F E

•

W A T E R

U.S. Environmental Protection Agency

Region III

Chesapeake Bay Program Office

Annapolis, Maryland
ACKNOWLEDGEMENTS

ii

Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

Technical Editorial Team and Contributing Authors

Multiple authors contributed to the report and an editorial team combined the various sections 

and also revised the report based on technical and Chesapeake Bay Program (CBP) partner 

reviews. 

The editorial team included Greg Allen (USEPA), Fred Pinkney (USFWS), Mike Focazio (USGS), 

Jamie Mitchell (Hampton Roads Sanitation District), and Scott Phillips (USGS). 

Contributing authors included: Vicki Blazer (USGS), Barnett Rattner (USGS), Ashlee Harvey 

(Chesapeake Research Consortium); Maria Garcia (USEPA); Bruce Pluta (USEPA); Diana Eignor 

(USEPA); and Mary Ann Ottinger (University of Maryland)

Peer and CBP Partner Reviewers 

The editorial team and authors thank the four independent peer reviewers: Lisa Saban 

(Windward Environmental LLC), Scott Ator (USGS), Patrick Phillips (USGS), and Ted Smith 

(USEPA). Many CBP partners also provided review comments on the report including 

representatives from Maryland, Virginia, Pennsylvania, West Virginia, Delaware, and Washington 

DC. Additionally, contacts from each jurisdiction listed above provided review and ensured 

accurate use of findings from their respective 2010 water quality assessment reports. The 

authors also thank Dale Simmons (USGS) for her editorial review.

Disclaimer: Any use of trade, firm, or product names is for descriptive purposes only and does not imply endorsement by 

the U.S. Government. 

Suggested Citation: US Environmental Protection Agency, US Geological Survey, US Fish and Wildlife Service, 2012 Toxic 

Contaminants in the Chesapeake Bay and its Watershed: Extent and Severity of Occurrence and Potential Biological 

Effects, USEPA Chesapeake Bay Program Office, Annapolis, MD, December, 2012, 175 pages.
TABLE OF CONTENTS

Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

iii

» Acknowledgements ii

» Executive Summary v

» 1.0 Introduction and Background 1

» 2.0 Extent and Severity of Contaminant Occurrence 13

2.1 Polychlorinated Biphenyls 13

 2.2 Dioxins and Furans 25

 2.3 Polycyclic Aromatic Hydrocarbons 31

 2.4 Petroleum Hydrocarbons 39

 2.5 Pesticides 44

 2.6 Pharmaceuticals 59

 2.7 Household and Personal Care Products 72

 2.8 Polybrominated Diphenyl Ether Flame Retardants 77

 2.9 Biogenic Hormones 81

 2.10 Metals and Metalloids 90

» 3.0 Responses of Fish to Cumulative and Interacting Stressors 104

» 4.0 Contaminant Exposure and Responses in Wildlife 113

» 5.0 Summary and Conclusions 116

» 6.0 References 126

» Appendix A: Supplementary tables 172
Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

v

EXECUTIVE SUMMARY EXECUTIVE

SUMMARY

Purpose of the Report

For many years, scientists and resource managers have recognized that exposure to toxic 

contaminants can result in adverse effects on biological resources within the Chesapeake 

Bay and its watershed. In 2010, the Chesapeake Bay Program (CBP), a Federal-jurisdictional 

partnership, reported that 72 percent of the Bay’s tidal-water segments are fully or partially 

impaired as a result of the presence of toxic contaminants. In some areas of the Bay watershed, 

fish-consumption advisories have been established as a result of concentrations of toxic 

contaminants. In recognition of these issues, the CBP developed the Toxics 2000 Strategy, in 

which commitments were made to prevent and reduce inputs of chemical contaminant and 

to eliminate toxic impacts on living resources that inhabit the Bay and its tributaries. Since 

2000, new concerns, such as intersex conditions in fish, have arisen. Although the causes are 

undetermined, there is increasing evidence that contaminant exposures may play a role. In 

2010, the President’s Chesapeake Bay Executive Order (EO 13508) Strategy directed Federal 

agencies to prepare a report summarizing information on the extent and severity of occurrence 

of toxic contamination in the Bay and its watershed. Findings in this report will be used by 

the CBP partnership to consider whether to adopt new goals for reducing inputs of toxic 

contaminants entering the Bay. This report also identifies research and monitoring gaps that 

could be considered to improve the understanding of the extent and severity occurrence of 

toxic contaminants in the Chesapeake Bay and its watershed.
Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

vi

EXECUTIVE SUMMARY

Approach

The findings in this report are based on a review of integrated water-quality assessment reports 

from the jurisdictions in the Bay watershed (Delaware, Maryland, New York, Pennsylvania, 

Virginia, West Virginia, and Washington, D.C.), Federal and State reports, and articles in scientific 

journals. The authors focused on summarizing results of studies conducted mostly since 2000 

and, in particular, the 2010 jurisdictional water-quality assessment reports were used to define 

the extent and severity of occurrence of the following contaminant groups:

» Polychlorinated Biphenyls (PCBs)

» Dioxins and Furans 

» Polycyclic Aromatic Hydrocarbons (PAHs) 

» Petroleum Hydrocarbons 

» Pesticides 

» Pharmaceuticals 

» Household and Personal Care Products 

» Polybrominated Diphenyl Ethers (PBDEs) 

» Biogenic Hormones 

» Metals and Metalloids

The approach used to characterize the extent and severity of occurrence of contaminant 

groups is described in detail in Chapter 1 of this report. Extent is characterized as “widespread”, 

“localized”, or “uncertain” depending on the amount of information acquired from readily 

available reports and peer-reviewed literature and whether the contaminant has been detected 

throughout the watershed or only in a limited number of subwatersheds. Severity, as defined in 

this report, is based entirely on the jurisdictions’ impairment determinations as identified in the 

integrated assessment reports. Contaminants that have caused impairments in many locations 

are considered to have widespread severity, contaminants associated with impairments in few 

locations are classified as having localized severity, and other contaminants or contaminant 

groups are identified as having uncertain severity. Where possible and appropriate, additional 

information such as peer-reviewed literature is included to provide perspective on potential 

severity, including evidence of adverse sublethal effects at environmentally relevant 

concentrations. 
Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

EXECUTIVE SUMMARY

vii

Conclusions about the Extent and Severity of 

Occurrence of Contaminant Groups 

Overall conclusions about the extent of occurrence of contaminant groups examined in this 

report are --

» Widespread extent: For polychlorinated biphenyls (PCBs), 

polycyclic aromatic hydrocarbons (PAHs), herbicides (primarily 

atrazine, simazine, metolachlor, and their degradation products), 

and mercury, available information indicates widespread extent 

of occurrence throughout the Bay watershed. 

» Localized extent: For dioxins/furans, petroleum 

hydrocarbons, some chlorinated insecticides (aldrin, chlordane, 

dieldrin, DDT/DDE, heptachlor epoxide, mirex), and some metals 

(aluminum, chromium, iron, lead, manganese, zinc), available 

information indicates localized extent of occurrence. 

» Uncertain extent: For pharmaceuticals, household and 

personal-care products, polybrominated diphenyl ether (PBDE) 

flame retardants, some pesticides, and biogenic hormones, 

available information is insufficient to determine extent of 

contamination. However, the widespread distribution of known 

sources of these contaminants (e.g., wastewater effluents, 

agricultural runoff, etc.) in the watershed and the summarized occurrence data indicate that 

some contaminants from each of these groups may have the potential to be found in many 

locations throughout the Bay watershed. 

Overall conclusions about the severity of contaminant groups examined in this report are --

» Widespread severity: For PCBs and mercury, impairments have been identified in many 

locations in the watershed, largely in response to concentrations in sediments and in fish 

tissues that frequently result in the need for fish-consumption advisories. 

» Localized severity: For dioxins/furans, PAHs, petroleum hydrocarbons, some chlorinated 

pesticides (aldrin, chlordane, dieldrin, DDT/DDE, heptachlor epoxide, mirex), and some 

metals (aluminum, chromium, iron, lead, manganese, zinc), the report identifies localized 

severity on the basis of impairments in a limited number of areas in the Bay watershed. 

esapeake Bay Program

00-YOUR-BAY

://www.epa.gov/r3chespk

The Chesapeake Bay Watershed

64,000 Square Miles of Land, Water, and People

"A Better Bay Through Better Science"

1997

Produced by the USGS from a mosaic of Landsat satellite imagery acquired from 1990-1994

U.S. Geological Survey 

 http://www.usgs.gov

1-888-ASK-USGS
Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

EXECUTIVE SUMMARY

viii

» Uncertain severity: For atrazine, some pharmaceuticals, some household and personal-care 

products, some PBDEs, and biogenic hormones, severity as defined in this report could not 

be assessed. However, recent peer-reviewed research has documented sublethal effects for 

some compounds at environmentally relevant concentrations, raising concerns about the 

potential for adverse ecological effects. 

Biological Effects of Toxic Contaminants on 

Fish and Wildlife

Additional supporting information on the toxic effects 

of contaminants on fish and wildlife is summarized 

to inform the discussion of severity. This information 

provides insights that can be used in assessing the 

cumulative and interacting effects of toxic chemicals 

as well as other stressors on fish and wildlife.

The following indicators of compromised fish health 

have been observed within populations in the 

Chesapeake Bay watershed: increased incidence 

of infectious disease and parasite infestations 

contributing to increased mortality in several species of fish; feminization (intersex, plasma 

vitellogenin) of largemouth and smallmouth bass and other signs of endocrine disruption; 

reduced reproductive success and recruitment of yellow perch in tributaries in certain highly 

urbanized drainage basins; and tumors in bottom-dwelling fish. The evidence for associations 

between exposure to toxic contaminants and these indicators of compromised fish health is 

discussed. 

Indications of responses to contaminant exposure have also been found among wildlife in the 

Chesapeake Bay watershed, primarily wild birds. In a few locations, eggshell thinning associated 

with p,p’-DDE is apparent, and reproduction may be impaired. In some cases, organochlorine 

pesticides are found in eggs of predatory birds at concentrations associated with embryo 

lethality. Several studies are cited in which PCB concentrations in addled bald eagle eggs may 

have been high enough to contribute to the failure to hatch. Detectable concentrations of 

PBDEs have been found in eggs of predatory birds that approach the lowest-observed-adverseeffect level for pipping and hatching success. 

Smallmouth bass with skin lesion. 
Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

EXECUTIVE SUMMARY

iv

Research and Monitoring Gaps

Monitoring gaps were identified for the following contaminant groups: dioxins and furans, 

petroleum hydrocarbons, some pesticides currently in use (e.g., insecticides and fungicides), 

pharmaceuticals, household and personal-care products, flame retardants, and biogenic 

hormones. Biological monitoring at many levels of biological organization (molecular to 

population) along with systematic evaluation of water contaminants and other stressors 

would allow for more effective documentation of the extent and severity of occurrence of toxic 

contaminants in the watershed.

Research gaps that limit understanding of the relations between sources of these contaminants, 

their pathways to the environment, and exposures to receptor organisms are identified. 

Research that accounts for the complexities of the effects of contaminant mixtures and multiple 

stressors, sublethal effects, nonlinear dose-response curves, and the role of contaminant 

exposure in immune response and subsequent pathogenic disease would help to define 

relations between contaminant exposures and potential effects in fish and wildlife. 
1.0 INTRODUCTION & BACKGROUND

1

Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

For many years, scientists and resource managers have recognized that exposure to toxic 

contaminants can result in adverse effects on biological resources within the Chesapeake 

Bay and its watershed. Some contaminant effects, such as those from chlorinated pesticides, 

are well documented and have been addressed through various approaches to minimize 

the occurrence of targeted contaminants. Other contaminants with known effects, such as 

polychlorinated biphenyls (PCBs), continue 

to enter the Bay’s ecological system. The 

potential for a range of land use activities such 

as human and animal waste management 

to provide sources of contaminants, such as 

pharmaceuticals, household and personal care 

products, and biogenic hormones has been 

documented; however, complete depictions of 

their occurrence, pathways to the environment, 

relative source contributions, and severity of 

effects from these environmental contaminants 

are the subjects of active research. The presence 

of toxic contaminants in the Chesapeake Bay has 

led to:

» the Chesapeake Bay Program (CBP) adopting the Toxics 2000 strategy (USEPA 2000a), which 

made commitments to prevent and reduce chemical contaminant inputs and eliminate 

toxic impacts on living resources that inhabit the Bay and its tributaries,

» impairment of the quality of living-resource conditions to the extent that 72 % of the 

Bay and its tidal river segments (2010) are fully or partially impaired as a result of toxic 

contaminants (Figure 1), 

» fish consumption advisories as a result of concentrations of certain toxic contaminants in 

fish in the Bay and its watershed, 

» research indicating that conventional toxicological benchmarking approaches may not 

adequately represent the potential for contaminants to do ecological harm, 

» realization that contaminants in the environment occur in mixtures that reflect complex 

combinations of land uses and contaminant sources and,

» the President’s Executive Order strategy for protecting and restoring the watershed calling 

for new reduction goals and strategies for toxic contaminants. 

Source: Jane Thomas/IAN Image Library
Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

2

1.0 INTRODUCTION & BACKGROUND

Figure 1 – Tidal segments with full or partial impairments due to toxic contaminants

The President’s Chesapeake Bay Protection and Restoration Executive Order (EO 13508) Strategy 

(May 12, 2010) directed Federal agencies to prepare a report summarizing information on the 

extent and severity of toxic contaminants in the Bay and its watershed. The findings in the 

report will be used by the CBP in 2013 to consider whether to establish new goals for reducing 

the input of toxic contaminants and, if established, to develop strategies by 2015 to carry out 

the goals. 
Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

3

1.0 INTRODUCTION & BACKGROUND

Progress on Previous Agreements (Toxics 2000)

During December 2000, the CBP Executive Council adopted “Toxics 2000 Strategy: A Chesapeake 

Bay Watershed Strategy for Chemical Contaminant Reduction, Prevention, and Assessment” (USEPA 

2000a). The agreement made substantial commitments to:

» prevent and reduce chemical contaminant inputs and eliminate toxic impacts on living 

resources that inhabit the Bay and rivers

» eliminate all chemical contaminant-related fish consumption bans and advisories

» clean up contaminants in the sediment in the three most urbanized areas referred to as 

“Regions of Concern” (i.e., Baltimore Harbor, Anacostia River, Elizabeth River)

» sustain progress in the face of increasing population and expanded development within the 

watershed. 

Since 2000, the reduction of nutrient and sediment inputs has been the main emphasis of CBP 

activities. Progress has been made by Federal and State agencies as well as non-government 

organizations (NGOs) that are completing ongoing work to control chemical contaminants. 

Federal agencies such as the United States Environmental Protection Agency (USEPA) have 

continued to oversee and conduct numerous contaminated site cleanups that have improved 

conditions in the Bay and in the watershed. The jurisdictions have continued to enforce permit 

conditions including industrial wastewater permits. The jurisdictions have also continued to 

monitor fish tissue and other environmental media to fulfill their data needs for determining 

fish consumption advisories and impairment listings. Federal and jurisdiction agencies charged 

with implementing and enforcing the hazardous material and waste statutes that control the 

release of toxic contaminants have continued to fulfill their obligations. Federal agencies with 

science-based missions such as U.S. Geological Survey (USGS), National Oceanographic and 

Atmospheric Administration (NOAA), U.S. Fish and Wildlife Service (USFWS) and USEPA monitor 

the presence of chemical contaminants and assess possible ecological effects. Many of the 

results of Federal monitoring efforts are discussed in this report. 

Progress has been made in at least two of the three previously designated Regions of Concern, 

the Elizabeth River and Anacostia River, due in part to the leadership provided by the Elizabeth 

River Project and Anacostia Watershed Restoration Partnership. For example in the Elizabeth 

River watershed, contaminated soil at a former naval shipyard was removed and the site was 

replanted to create a wetland. Multiple industrial sites are being cleaned up in the Elizabeth 
Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

4

1.0 INTRODUCTION & BACKGROUND

River to reduce bottom sediment contaminated with PAHs and other pollutants. In the 

Anacostia watershed, stormwater retrofit projects have been completed to allow for improved 

treatment of stormwater originating from hundreds of acres in the river’s watershed. The 

Anacostia is benefiting from a Total Maximum Daily Load (TMDL) that targets trash, which 

will reduce inputs of contaminants associated with household products and other industrial 

sources of waste. The Anacostia Watershed Restoration Plan (Anacostia Watershed Restoration 

Partnership 2010) is being implemented through multijurisdictional cooperation and includes 

projects that will reduce inputs of toxic contaminants to the river. During 2012, the USEPA 

Chesapeake Bay Program Office focused one million dollars of grant funds toward the Anacostia 

watershed. Both the Anacostia watershed and Baltimore Harbor were chosen for USEPA’s Urban 

Waters Initiative, which is working to align Federal programs and investments and build local 

capacity for improving ecological conditions in these watersheds.

In 2006, the CBP completed an analysis of information that led to prioritization of organic 

pollutants for use in developing management strategies for reducing pollutant inputs. 

Although several of the contaminant groups that were identified as high priority in 2006 are 

also identified in this report, the 2006 prioritization was not substantially referred to in this 

report because the project team believed more current information was available. Strategies for 

reduction of those high priority pollutants were in development when the CBP organizational 

decision was made in 2007, to disband the former CBP Toxics Subcommittee to allow for greater 

focus on development of the nutrient and sediment TMDL. Prior to 2007, the efforts of the 

Toxics Subcommittee focused on further characterizing the condition of the Bay with regard to 

ecological impacts from toxic contaminants. The contaminants characterization data generated 

during that time are referred to in this report. 

It is beyond the scope of this report to make a quantitative assessment of the progress made 

on the original commitments in the Toxics 2000 strategy. Since the strategy was written, the 

conditions that existed remain. According to the environmental indicator maintained by 

the CBP (see Figure 1), which measures the number of tidal segments with a partial or full 

jurisdiction-listed impairment due to toxic contaminants in 2010, a similar extent of impairment 

exists in the Bay compared with the previous version, based on 2008 jurisdiction impairment 

listings. Research has augmented our understanding of sublethal effects of contaminant 

mixtures and new issues, such as intersex characteristics in fish in the Bay watershed, have 

arisen. The focus of this report, therefore, is to summarize the current conditions of extent and 
Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

5

1.0 INTRODUCTION & BACKGROUND

1. Polychlorinated Biphenyls (PCBs) 6. Pharmaceuticals

2. Dioxins and Furans 7. Household and Personal Care Products

3. Polycyclic Aromatic Hydrocarbons (PAHs) 8. Polybrominated Diphenyl Ether (PBDE) Flame Retardants

4. Petroleum Hydrocarbons 9. Biogenic Hormones

5. Pesticides 10. Metals and Metalloids

severity of effects from toxic contaminants. The report findings will be used to assist the CBP in 

considering goals and strategies to reduce risk to the Bay’s biological resources.

Report Purpose and Scope 

This report summarizes readily available and acquired information about the extent of 

occurrence and severity of effects for the following groups of toxic contaminants in Chesapeake 

Bay and its watershed (Table 1). 

Table 1 List of Contaminant Groups

The report also provides considerations for developing reduction goals if established, and 

identifies research and monitoring that could be conducted to better define the extent and 

severity of groups of contaminants. The report focuses on the severity of adverse effects of 

toxic contaminants on natural resources in the Bay and its watershed. It does not address 

potential effects on human health except in recognizing fish impairments and the status of 

fish consumption advisories established by the jurisdictions in the watershed. The extent of 

occurrence of toxic contaminants is defined for major groups of contaminants, such as PCBs 

and pesticides, which are known to occur in the Bay and its watershed. Information from the 

jurisdictions’ integrated water quality assessment reports is a key resource in helping to define 

the extent and severity of toxic contaminants. Additional information from previously published 

Federal and academic studies was examined. For some groups of contaminants, such as PCBs, 

polycyclic aromatic hydrocarbons (PAHs), and some pesticides and metals, available information 

was used to characterize the contaminant as widespread, localized, or uncertain. For other 

groups of contaminants, including pharmaceuticals, personal care products, flame retardants, 

and hormones, data were limited; therefore, conclusions about extent of occurrence are 

constrained and less certain. 
Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

6

1.0 INTRODUCTION & BACKGROUND

The findings will be used by the CBP to consider whether to establish new or updated goals for 

reduction of toxic contaminants. The primary audience is the decision makers in the CBP who 

are working to manage fisheries, habitat, water quality, and healthy watersheds. Several CBP Goal 

Implementation Teams (GITs) will be informed by this report on the effects of toxic contaminants. 

The Fisheries GIT may use the findings to better understand the health of fisheries in the Bay and 

its watershed. The Habitat GIT may use the findings to understand effects on wildlife (especially 

waterfowl) that use coastal wetlands and submerged aquatic vegetation. Because the Water 

Quality GIT is working to make waters both fishable and swimmable, the information developed 

in this report may help to develop nutrient and sediment reduction efforts that are designed 

with consideration of potential toxic contaminant effects. The Healthy Watersheds GIT may work 

to prevent impacts of toxic contaminants on healthy watersheds. The Water Quality GIT will 

coordinate with the other GITs and will use the information to work with CBP leadership groups 

such as the Management Board and Principal’s Staff Committee to consider whether to establish 

new goals to reduce toxic contaminants (during 2013) and, if established, to develop more 

detailed strategies to carry out the goals (by 2015).

Assessment Approach

Studies published from 2000 onward were used as the source of current environmental data. 

The primary sources of information were studies prepared by scientists from State and Federal 

agencies, colleges and universities, consulting firms, and NGOs. It was beyond the scope of 

this report to assemble a toxic contaminant data base for the Chesapeake Bay watershed and 

statistically analyze and interpret that data. Trend analysis was also beyond the scope of this report. 

Chapter 2 of the report is a detailed technical assessment of the extent and severity of toxic 

contaminant concentrations and effects for ten contaminant groups (Table 1). These contaminant 

groups include both chemical classes (e.g., PCBs and PAHs), products with similar use (e.g., 

pharmaceuticals), and biological products (biogenic hormones). The contaminant groups were 

selected as representative of major categories of contaminants in terms of potential natural 

and anthropogenic sources and pathways to the environment, mobility in the environment 

and potential for widespread extent, and known or suspected adverse ecological impact. Thus, 

the selected contaminant groups do not include all potential toxic contaminants that could be 

found in the watershed. For example, volatile organic compounds (VOCs) were not assessed as 

a specific group of compounds but many individual VOC compounds were included within the 

representative contaminant groups. These determinations were based on best professional 
Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

7

1.0 INTRODUCTION & BACKGROUND

judgment of the writing team and include chemicals that have been studied for decades as well 

as contaminants of emerging concern. Within each section of Chapter 2, representative and 

readily available concentrations of the contaminant groups in various media (usually surface 

water, sediment, and tissue from fish, birds, and mammals) are summarized.

A consistent approach (Figure 2) is used to evaluate the extent and severity of toxic chemical 

effects in the Bay watershed for each of the ten contaminant groups. Representative and readily 

available information (i.e. jurisdiction assessment reports, summaries of databases, published 

reports and journal articles) was acquired and considered for each contaminant group. 

Geographic extent was categorized as “widespread” if detectable concentrations were widely 

distributed across the Chesapeake Bay watershed and “localized” if only in specific areas. This 

is based on representative environmental media including water, sediment, and biota tissue as 

acquired. In cases where occurrence data were not available, other factors known or suspected 

to control the geographic extent of a contaminant or contaminant group, such as sources, land 

use, and pathways to the environment are considered and discussed.

The assessment of severity was based primarily on the monitoring reports that states are 

required to prepare under the Clean Water Act. Section 305(b) requires each State (and the 

District of Columbia) to monitor, assess and report on the quality of its waters in terms of 

designated uses. These uses include supporting aquatic life, fish consumption, recreation, 

and shellfish harvesting. Monitoring and assessment data are evaluated with respect to these 

designated uses. If the state agency interprets the concentration data as exceeding the state 

water quality standards (e.g., due to water column concentrations of specific contaminants or 

fish tissue concentrations that limit human consumption), the state identifies the water body 

as impaired for a particular designated use. Section 303(d) of the Clean Water Act requires 

each State to compile a list (Impaired Waters List) identifying those waters not meeting water 

quality standards. Impaired waters are defined as any water bodies that are not supporting 

one or more designated uses. Every two years, each State either prepares a report for each 

section of the Clean Water Act or an Integrated Water Quality Report that covers both sections. 

These reports describe the status of all assessed waters, list impaired waters and the causes of 

impairment, and provide the status of actions being taken to restore impaired waters. 
Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

8

1.0 INTRODUCTION & BACKGROUND

Causes of impairment include chemical contaminants (such as PCBs and metals), physical 

conditions (such as elevated temperature, excessive siltation, or alterations of habitat), and 

biological contaminants (such as bacteria).

This report used the latest available (2010) water quality reports prepared by the jurisdictions 

in the Bay watershed (Delaware, Maryland, New York, Pennsylvania, Virginia, West Virginia, 

and Washington, D.C.) as the primary sources of information for describing the extent of 

impaired waters within the Bay watershed. The agencies include: Delaware Department 

of Natural Resources (DNREC), Maryland Department of the Environment (MDE), New York 

State Department of Environmental Conservation (NYSDEC), Pennsylvania Department of 

Environmental Protection (PADEP), Virginia Department of Environmental Quality (VADEQ), West 

Virginia Department of Environmental Protection (WVDEP), District of Columbia Department 

of the Environment (DDOE) The report focuses on impairments due to chemical contamination 

that limit aquatic life support and/or limit fish consumption. Within each State, only those subwatersheds that are part of the Chesapeake Bay watershed were evaluated. If impairments 

are identified at many locations in the watershed, severity is classified as “widespread”; if 

impairments are identified at few locations, severity is classified as “localized” (Figure 2). The 

uncertainties associated with these categories are discussed below. 

A second closely related major source of information on severity of a contaminant group was the 

jurisdictions’ listings of fish tissue advisories, which overlaps considerably with the waters listed 

as impaired for fish consumption. The advisories were screened against the Chesapeake Bay 

watershed boundaries. Although this report is largely focused on ecological rather than human 

health effects of contaminants, restrictions on fish consumption are included because they 

represent a lost or restricted use of the water resource.

A third major source of information for describing severity was the sediment chemistry, benthic 

macroinvertebrate community, and sediment toxicity studies conducted at various locations 

within the watershed. The synoptic collection of these three types of data (using the same 

sampling locations and timing) is termed the “sediment triad approach” (e.g., Chapman 1990) and 

is a useful method for characterizing the quality of freshwater, estuarine, and marine habitats. 

A Bay-wide study of 210 sediment sampling locations was conducted by the National Oceanic 

and Atmospheric Administration (NOAA) in 1998, 1999, and 2001 and reported by Hartwell and 

Hameedi (2007). Smaller scale sediment triad studies within the Bay watershed were prepared by 

Pinkney et al. (2005) and Fulton et al. (2007).
Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

9

1.0 INTRODUCTION & BACKGROUND

Figure 2. Approach for assessing the extent and severity of effects for each 

contaminant group within the Chesapeake Bay watershed. 

 Acquire readily-available reports and peer-reviewed literature for each contaminant group.

 Possible outcomes:

 1) Limited information within the Chesapeake Bay watershed 

 2) Extensive information throughout the Chesapeake Bay watershed

Summarize the acquired information on the geographic extent of the contaminant group 

 within the Chesapeake Bay watershed.

 Possible outcomes:

 1) Localized: Detectable concentrations in a limited number of subwatersheds

 2) Widespread: Detectable concentrations are found throughout the Bay watershed

 3) Uncertain

 Summarize the acquired information on the severity of contaminant effects within the 

 Chesapeake Bay watershed.

 Possible outcomes:

 1) Localized: Impairments in a few locations

 2) Widespread: Impairments at many locations

 3) Uncertain

Additional reports on Bay tributaries or watersheds aimed at specific environmental questions or 

problems were summarized. Examples include a series of studies of fish kills within the Potomac 

watershed (e.g., Blazer et al. 2007, 2010; Ciparis et al. 2012); fish tumor surveys (e.g., Pinkney et 

al. 2009; Vogelbein and Unger 2006); and assessments of the status of the Anacostia River (e.g., 

Anacostia Watershed Toxics Alliance 2009, Velinsky et al. 2011; McGee et al. 2009). 

Benchmarks, defined as standards or points of reference used for comparisons or assessments, 

were identified and used to provide context for the concentration data. For many contaminant 

groups, however, no benchmarks are available. Study authors, including state agencies, 

frequently compare water column concentrations with state water quality standards (Appendix 

A, Table A-1). Since the Bay states do not have sediment quality standards, the study authors 
Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

10

1.0 INTRODUCTION & BACKGROUND

and state agencies compared sediment concentrations with guidance values they identified 

(Appendix A, Table A-2). The most commonly used guidelines for evaluating the severity of 

sediment contamination in estuarine and marine sediment are provided by the effects rangelow (ERL) and effects range- median (ERM) developed by NOAA (Long et al. 1995). The ERM is 

the 50th percentile concentration of a data set reporting adverse effects to a variety of benthic 

invertebrates from estuarine and marine environments; adverse effects occur more frequently 

than not above this concentration. The ERL is a lower threshold concentration (10th percentile 

value from the same data set), below which toxic impacts are unlikely to occur; concentrations 

between the ERL and ERM may occasionally result in adverse effect (Long et al. 1995). For 

freshwater sediments, consensus-based guidance values were derived as threshold effects 

concentrations (TECs) and the probable effects concentration (PECs) (MacDonald et al. 2000) 

and are functionally similar to the ERL and ERM. TECs identify contaminant concentrations below 

which harmful effects on sediment dwelling organisms are not expected. PECs are contaminant 

concentrations above which harmful effects on sediment-dwelling organisms are expected to 

occur frequently. 

All four sediment guidance values are empirically based and are used in this report as screening 

values and not as risk thresholds. Other more theoretically based guidance values include the 

equilibrium partitioning approach and evaluation of the relationship between simultaneously 

extracted metal concentrations (SEM) and acid volatile sulfide (AVS) concentrations. In the 

reports summarized, the authors compared sediment contaminant concentrations with ERL, 

ERM, TEC, and PEC values. In the current report, the higher benchmarks (ERM and PEC), which 

are more frequently associated with adverse effects on the benthic community than the lower 

benchmarks (ERL and TEC), are emphasized. The authors relied on the guidance values used in 

the cited studies because they did not have the resources to reevaluate the data and reinterpret 

the results using alternate approaches. It is recognized that these thresholds are only based 

on direct effects on benthic organisms and do not address effects on fish (e.g., development of 

liver pathology such as tumors) or concerns about bioaccumulation. There are also limitations 

inherent in relying on these empirically-based guidance values, especially for mixtures of PAHs 

where the narcotic mode of action provides a theoretical basis for applying the equilibrium 

partitioning approach to estimate benchmarks (DiToro and McGrath 2000). 
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1.0 INTRODUCTION & BACKGROUND

For tissue data (largely for fish, birds, and mammals), the residue concentrations were compared 

against critical residue thresholds (those associated with toxic effects) but again, many of the 

contaminants assessed have limited, or no, threshold values. Many thresholds have been 

compiled and described in three books (Beyer et al. 1996, Hoffman et al. 2003, and Beyer and 

Meador 2011), which were used as the primary source of this information in the report. Food 

chain modeling to assess risks to birds and mammals was beyond the scope of the report. 

Chapter 3, Responses of Fish to Cumulative and Interacting Stressors, summarizes studies that 

indicate the adverse effects of the exposure of fish to complex mixtures of both traditional 

contaminants, contaminants of emerging concern, and multiple stressors (both biological and 

chemical). Chapter 4, Contaminant Exposure and Responses in Wildlife, is a synopsis of the wildlife 

information spread among the sections on individual contaminant groups in Chapter 2, and is 

largely focused on effects on birds and mammals. A major source of the wildlife data was the 

Contaminant Exposure and Effects—Terrestrial Vertebrates (CEE-TV) database (www.pwrc.usgs.

gov/contaminants-online) containing over 20,000 geo-referenced data records for marine and 

estuarine habitats across the nation (Rattner et al. 2005). Chapter 5 presents the summary and 

conclusions of the report.

There are uncertainties involved in summarizing and categorizing the extent and severity of 

toxic chemical effects for any contaminant or contaminant group. All determinations and the 

use of the words “limited”, “extensive”, “localized”, and “widespread” reflect the lack of precise 

decision boundaries; however, the professional judgment used to support each of these 

determinations is supplied. Lack of sufficient geographic coverage of monitoring data across 

the watershed often prevents conclusions on widespread versus localized or uncertain extent 

of occurrence. In some cases, determining that there is widespread as opposed to localized 

severity was readily apparent where there are impairments identified at many locations in 

the watershed. It was beyond the scope of this report to make spatially explicit delineations 

(such as percentage of stream miles impaired) to establish rules for distinguishing localized 

versus widespread severity. Additional uncertainty in this report is inherent in its reliance on 

the jurisdictions’ water quality reports to describe the severity of contaminant impacts. The 

jurisdictions have different methods of summarizing their monitoring data and comparing with 

standards, which may vary, to determine whether a water body is impaired. For many of the 

chemicals of emerging concern, there are no existing benchmarks or standards, and, therefore, 

no opportunity for the jurisdictions to identify impairments.
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1.0 INTRODUCTION & BACKGROUND

There may be substantial differences in detection limits used in different studies. In some 

cases, detection limits for certain analytes were above concentrations associated with adverse 

effects or even above state water quality standards. These and other factors can lead to 

additional uncertainty regarding the geographic extent and/or severity of a contaminant 

group. Consequently in some cases, a determination of the extent or severity was unsupported 

and a determination of “uncertain” was made (Figure 2). An uncertain determination does not 

imply presence or absence of potential extent or severity; where possible, relevant exposure 

and toxicity information was summarized to inform readers on the current status of these 

contaminant groups. 

Finally, recent literature on sublethal effects of contaminants within multiple stressor 

environments has drawn attention to the potential inadequacies of conventional contaminantby-contaminant benchmarking approaches (Feingold et al. 2010; Burton et al. 2012). Full 

consideration of this rapidly emerging aspect of environmental toxicology was not possible 

for determinations of severity as defined in this report. In some cases, the authors have 

provided additional context for considerations of severity of contaminants and contaminant 

groups by citing recent relevant peer-reviewed literature on the potential ecological effects of 

these contaminants. Representative toxicity studies conducted at environmentally relevant 

concentrations (i.e. concentrations known to occur in the watershed) were summarized. 

However, frequently there is insufficient monitoring data to determine the duration of exposure 

of sensitive organisms or vulnerable life stages, and the likelihood of adverse effects in the field. 
2.0 EXTENT AND SEVERITY OF CONTAMINANT OCCURRENCE

13

Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

The extent and severity of occurrence for the 10 groups of contaminants listed in Table 1 are 

presented in the following subsections. There is a subsection for each group, which includes an 

abstract to summarize the main points and conclusions.

2.1 Polychlorinated Biphenyls

Abstract

PCBs are a group of synthetic organochlorine chemicals widely used as dielectric and coolant 

fluids in transformers and capacitors. In 1977, the U.S. banned the production of PCBs out of 

concern for their persistence in the environment as well as evidence indicating that PCBs were 

bioaccumulative and has the potential to cause toxic impacts. Though PCBs are no longer 

produced, there are continued authorized uses of PCB-containing materials which pose the 

potential for environmental release. The inadvertent production of PCBs in certain manufacturing 

processes represents an additional contemporary source. Impairments resulting from PCB 

contamination are widespread. The Commonwealth of Virginia documents water column 

impairments in the Potomac and Shenandoah Rivers impacting nine river miles and one square 

mile of estuarine waters. All of the Bay jurisdictions have waterbodies identified as impaired 

for human consumption of fish tissue. The District of Columbia and the State of New York have 

general fish consumption advisories out of concern for contamination of fish tissue from PCBs 

and other chemical contaminants. The State of New York includes a fish tissue impairment for 

Koppers Pond, located in the Chemung River basin. Delaware identifies a fish tissue impairment 

for the Chesapeake and Delaware Canal (C&D Canal). Maryland lists 30 assessment segments 

as impaired for fish tissue consumption. In Pennsylvania, the Susquehanna River is impaired 

for 208 river miles, extending from the 

New York/Pennsylvania state line to the 

city of Sunbury, PA. Virginia lists five Bay 

tributaries as impaired, impacting 456 river 

miles and 2,011 square miles of estuarine 

waters. In West Virginia, two waterbodies 

are impaired by PCB contamination in fish 

tissue: the South Branch of the Potomac 

River and the Shenandoah River. Unlike 

concentrations of chlorinated pesticides, 

concentrations of PCBs in tissues of many 

species of Chesapeake Bay wildlife have 
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2.0 EXTENT AND SEVERITY OF CONTAMINANT OCCURRENCE

not declined since the final USEPA rule restricting the manufacture, processing, and distribution 

of these compounds became effective in 1979. The available data indicates that the extent of PCB 

contamination within the Chesapeake Bay watershed is widespread, being detected in tissue and 

sediment in most sampling locations. 

 

Background

PCBs are a group of synthetic organic chemicals with high thermal stability, making them 

important in applications such as dielectric fluids in transformers and capacitors, heat transfer 

fluids, and lubricants. In addition, PCBs were used in plasticizers (e.g., carbonless paper), 

inks, adhesives, sealants and caulk. There are no natural sources of PCBs to the environment. 

PCBs typically exist as mixtures of chlorinated biphenyl compounds with varying degrees of 

chlorination. A total of 209 possible compounds, known as congeners, result from the variation of 

chlorination (1 – 10 chlorine atoms) around the biphenyl rings (Agency for Toxic Substances and 

Disease Registry (ATSDR) 2000). 

PCBs are relatively insoluble in water with solubility decreasing with increasing chlorination. 

These hydrophobic compounds dissolve readily in nonpolar organic solvents and in biological 

lipids. Due to different degrees of chlorination, the physical and chemical properties vary among 

the congeners (ATSDR 2000).

PCBs have not been produced in the U.S. since August of 1977 due to evidence that this group 

of compounds was persistent and bioaccumulative in the environment and had the potential 

to cause toxic effects. Aside from the historical contributions of PCBs to the environment, PCBs 

continue to be released to the environment, however, through leaks or fires in PCB-containing 

equipment, accidental spills, illegal or improper disposal, burning of PCB-containing oils in 

incinerators, and leaks from hazardous waste sites (Total Maximum Daily Load reports: MDE 

2009a, 2009b, 2009c, 2011a, 2011b, USEPA and VADEQ 2001, Haywood and Buchanan 2007). 

Point source discharges are regulated; however, discharges of PCBs may continue as a result 

of historical contamination or inadvertent production (Oregon Department of Environmental 

Quality (ODEQ) 2012, Du et al. 2008, Hu and Hornbuckle 2010). Specific processes implicated in 

inadvertent production have been identified as those that involve chlorinated solvents, paints, 

printing inks, agricultural chemicals, plastics and detergent bars (ODEQ 2012). 
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2.0 EXTENT AND SEVERITY OF CONTAMINANT OCCURRENCE

In addition, there are continued authorized uses of PCBs including in sealed systems such as 

transformers and some heat transfer systems (ODEQ 2012). ODEQ (2012) provides additional 

information on potential sources in waste materials and recycling operations. 

In the aqueous environment, the higher molecular weight PCBs (i.e., more chlorinated) are 

typically sorbed to suspended solids and sediment whereas the lower molecular weight PCBs 

tend to volatilize to the atmosphere. Once in the environment, PCBs cycle among environmental 

media (air, water, soil/sediment, biota). Volatilized PCBs are redeposited to land and water through 

precipitation events (ATSDR 2000). 

PCBs in the water column can be removed through volatilization at the air-water interface, 

through sorption to sediments and suspended solids, and by uptake in aquatic organisms (ATSDR 

2000). Uptake in aquatic organisms can occur through bioconcentration and/or bioaccumulation. 

In bioconcentration, uptake occurs directly from the water column whereas bioaccumulation 

occurs through the combined uptake of food, water and sediment. Concentrations of PCBs 

increase through the higher trophic levels. As a result of the lipophilicity of these compounds, 

they tend to accumulate within the tissues of the organisms (ATSDR 2000). 

PCBs are listed as probable human carcinogens by the USEPA and the International Agency 

for Research on Cancer (ATSDR 2000). Studies of workers reported that PCBs were associated 

with cancer of the liver and biliary tract. ATSDR (2000) summarized studies that reported that 

women who consumed high amounts of PCB-contaminated fish gave birth to babies with 

lower birth weights. ATSDR noted that infants of these women had abnormal responses to 

behavior tests and that the infants’ problems with motor skills and short-term memory persisted. 

Therefore, prenatal exposure and exposure of children through breast milk is a concern, and fish 

consumption advisories for PCBs are therefore more restrictive for children and women of childbearing age than for the general population. 

The data for this chapter were derived primarily from the state water quality assessment reports 

(DDOE 2010, MDE 2010, NYSDEC 2007, PADEP 2010, VADEQ 2010, WVDEP 2010) which documents 

a water body’s attainment of its designated use. Water bodies that fail to meet the water quality 

standards or criteria applicable for the state’s designated use are categorized as “impaired”. These 

numerical thresholds may differ from state to state. Comparisons between the state standards 

will not be made; however, impairment identifications are noted. For informational purposes, 
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2.0 EXTENT AND SEVERITY OF CONTAMINANT OCCURRENCE

current jurisdiction water quality standards for aquatic life protection and screening values for 

sediment quality are located in the appendices.

Though the integrated assessment reports do not provide quantitative data on PCB 

concentrations, such data were available from the TMDL reports prepared for several impaired 

waterways in the Chesapeake Bay watershed. Under the Clean Water Act, states must develop 

TMDL reports for each impaired water body, identifying the probable sources of impairment and 

the required load reductions from each source category necessary to comply with the standard. 

Several TMDL reports in the Chesapeake Bay watershed have been completed for fish tissue 

impairments and provide a source for quantitative data. NOAA provided an additional source 

of quantitative data in its sediment survey of the Chesapeake Bay conducted during 1998 to 

2001, characterizing PCB concentrations across a large portion of the watershed (Hartwell and 

Hameedi 2007). 

Measuring low concentrations of PCBs is subject to analytical challenges. Though not yet 

promulgated in the Federal Register for Clean Water Act programs, USEPA Method 1668, a lowlevel PCB method, is being used to support many TMDL studies. Though there is on-going debate 

and concern about the method’s reliability at concentrations near its reported detection limit 

and its sensitivity to false positives, USEPA indicated that its use in state TMDL programs has been 

successful (USEPA 2012b). 

Water

There are two categories of water quality standards applicable to PCBs: standards developed for 

the protection of aquatic life and standards developed for the protection of human health from 

the consumption of fish. Documented water column exceedances of the state water quality 

standards are uncommon. This could be attributable, in part, to the limitations of the analytical 

methods most commonly used for the routine assessment of state waters. Routine methods 

typically quantify a small subset of the total 209 PCB congeners and have analytical detection 

limits several orders of magnitude greater than the state standards (USEPA 2012a). 

The Commonwealth of Virginia documents water column impairments in the Potomac and 

Shenandoah river basins. Approximately 9 river miles are impaired and 1 square mile of estuary 

is impaired as a result of PCB contamination (Virginia Department of Environmental Quality 

(VADEQ) 2010). 
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2.0 EXTENT AND SEVERITY OF CONTAMINANT OCCURRENCE

The multi-jurisdictional Potomac/Anacostia Rivers TMDL report identified a range of water column 

values for total PCBs from below detection to 0.34 μg/L (congener specific detection limits 2 - 8 

pg/L) (Haywood and Buchanan 2007). Completed Maryland TMDL studies reported water column 

values for total PCBs ranging from 0.00009 to 0.03071 μg/L (MDE 2011a, MDE 2009a, MDE 2009b, 

MDE 2009c, MDE 2009d). Virginia and West Virginia collaborated for the Shenandoah River TMDL 

and measured water column values for total PCBs ranging from 0.0000077 to 0.0000791 μg/L 

(USEPA 2001). Pennsylvania’s Susquehanna River TMDL report indicated a water column value of 

0.0276 μg/L for total PCBs (PADEP 1999). Water quality standards for total PCBs in water for these 

states range from 0.014 to 0.030 μg/L (freshwater and saltwater chronic values, respectively). 

 

Sediment

Maryland documents impairments based on exceedances of screening values in sediment 

(Bear Creek, Curtis Bay, and Baltimore Harbor, MDE 2010). The other states in the Chesapeake 

Bay watershed did not document impairments for PCBs in sediment. In Virginia, 208 stations 

were monitored for PCBs in sediment in conjunction with the Commonwealth’s freshwater 

probabilistic monitoring program. PCBs were detected in all samples but were below the PEC 

screening value of 676 ppb (VADEQ 2010). Virginia conducted additional sediment monitoring 

as part of a toxicological characterization effort (Roberts et al 2002, 2003, 2004). There were 

very few instances in which PCB congeners were detected in the sediment. In the Mattaponi 

and Pamunkey Rivers, the concentrations were “well below the concentrations expected to 

produce adverse impacts” (Roberts et al. 2004). In the Mattaponi, the results ranged from 

< 6 – 125 ppb dw and in the Pamunkey, the results ranged from < 5 – 70 ppb. Historical data 

from the Nanticoke River in Delaware indicate that PCB concentrations “were not detected at 

levels expected to pose a significant risk to aquatic life or human health” (DNREC 1997). More 

recent sediment core data were collected to support a maintenance dredging project in the 

Nanticoke (EA 2006). The resulting data showed that concentrations in the sediment ranged 

from approximately 29 – 44.5 ppb.

NOAA completed a survey of toxic contaminants in sediments Bay-wide from 1998 to 2001, 

including PCB analysis for a list of approximately 22 congeners. Sediment concentrations from 

samples in the Bay tributaries tended to be higher than those collected from the embayments 

and mainstem (Hartwell and Hameedi 2007) and ranged from below detection to 122 ppb. 
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2.0 EXTENT AND SEVERITY OF CONTAMINANT OCCURRENCE

None of the sediment concentrations of total PCBs exceed the ERM level of 180 ppb, however, 

several sites did document levels above the ERL of 22.7 ppb (Susquehanna Flats, Bay Bridge, and 

the Elizabeth River). The sediment values for the Chesapeake Bay embayments were all below 

the ERL. 

In 1994, NOAA released a report documenting sediment contamination in the Chesapeake 

and Delaware Bays (NOAA 1994). At that time, sediment concentrations of total PCBs around 

Fort McHenry (Baltimore, MD) ranked among the highest total PCB sediment concentrations 

documented at NOAA’s National Status and Trends (NS&T) sites around the country (above the 

89th percentile). The measured concentration of total PCBs at this site was 679 ppb, above the 

ERM of 180 ppb. The more recent NOAA report (Hartwell and Hameedi 2007) did not include data 

from the Fort McHenry area. 

The Potomac/Anacostia Rivers TMDL identified a range of sediment values of total PCBs from 

nondetect – 1,550 ppb dw (Haywood and Buchanan 2007). Completed Maryland TMDL studies 

reported sediment values ranging from 1.4 – 59.14 ppb dw (MDE 2011a, MDE 2009a, MDE 2009b, 

MDE 2009c, MDE 2009d). Virginia and West Virginia collaborated for the Shenandoah River TMDL 

study and measured sediment values for total PCBs ranging from 0.31 – 100 ppb (USEPA 2001). 

PCBs in sediments at concentrations above certain thresholds pose risks to aquatic life through 

several pathways. First, sediment-bound PCBs serve as a source for bioaccumulation in prey 

that ultimately results in fish contamination. This topic is covered extensively by Haywood 

and Buchanan (2007) in the TMDL document for the tidal Potomac River. They calculated 

bioaccumulation factor (BAF)-based target sediment concentrations of 2.8 to 12.0 ppb. These 

would translate to fish tissue concentrations at or below the impairment thresholds of the District 

of Columbia, Maryland, and Virginia. A high percentage of the sediments monitored in the tidal 

Potomac have total PCB concentrations above these targets, hence the requirement for a 96% 

reduction in PCB loading in order to achieve the TMDL (Haywood and Buchanan 2007). 

Eggs, larvae, and juveniles of Bay fish species are exposed to PCB-contaminated sediments and it 

is likely that these life stages are more sensitive than adults (Eisler and Belisle 1996). In addition, 

maternal transfer of PCBs occurs during oogenesis (Fisk and Johnston 1998). Calculations of 

sediment thresholds for toxic effects in fish have been conducted for juvenile salmonids by 

Meador et al. (2002) who evaluated 15 studies that reported total PCB tissue concentrations 
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2.0 EXTENT AND SEVERITY OF CONTAMINANT OCCURRENCE

and toxic effects. Using literature-based biota sediment accumulation factors (BSAFs) and lipid 

calculations, they suggested a sediment effect threshold concentration ranging from 75 to 600 

ppb dw. The range depended on the BSAF and percent total organic carbon in the sediments. 

These sediment thresholds for the protection of salmonids are higher than the TMDL targets of 

Haywood and Buchanan (2007).

Fish and Shellfish

Fish tissue advisories and impairments

The state water quality standards regulating fish tissue concentrations are designed to protect 

human health by minimizing dietary exposure to PCBs through fish consumption. These 

concentrations can be elevated to unacceptable levels as a result of interactions with sediment, 

the water column and through trophic transfer. Though high resolution data may not often 

be available for sediment and water in waterways that have fish tissue impairments, the 

bioaccumulation of PCBs in excess of state standards or screening values may indicate that total 

PCB concentrations in sediment and water are contributing to the impairment. 

All of the Bay jurisdictions have water bodies listed with fish consumption advisories due to 

PCB fish tissue concentrations in excess of a state standard or health department threshold. 

Most advisories limit exposure in terms of meals per week or month and there is variation in the 

formulas used to calculate the restrictions.

The District of Columbia issued fishing advisories for all its waters in order to minimize the risk of 

human exposure to elevated levels of PCBs and other chemicals in fish tissue (DDOE 2010). 

PCB contamination is the principal toxicological driver for the fish tissue impairment identified 

in Delaware’s Chesapeake and Delaware Canal (C&D Canal) (DNREC 2010 and R. Greene, DNREC, 

personal communication). The advisory is based on data generated for a 1999 report. Delaware is 

updating its toxics data for the C&D Canal in 2013 (R. Greene, DNREC, personal communication). 

The State of Maryland lists more than 30 segments in the Bay watershed for fish tissue 

impairment. In the Severn River mesohaline segment, the State’s assessment report indicates that 

fish tissue concentrations may be low enough to meet the standard; however, additional data are 

needed for confirmation (MDE 2010). Completed Maryland TMDLs document a range of fish tissue 

values of 22.1 to 608.9 ppb (MDE 2009a-d, MDE 2011a).
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2.0 EXTENT AND SEVERITY OF CONTAMINANT OCCURRENCE

The State of New York has a general fish consumption advisory for all state waters in order to 

minimize the risk of human exposure to elevated levels of PCBs and other chemicals in fish tissue. 

The New York Department of Health (NYDH) advises the public to consume no more than four 

one-half pound meals a month of fish, with no more than one meal per week (NYDH 2012). In the 

Chemung River basin, one impoundment is identified as impaired for PCBs in fish tissue (Koppers 

Pond) (NYSDEC 2007). 

Within the Chesapeake Bay watershed of Pennsylvania, the Susquehanna River has a PCB fish 

consumption advisory for 208 miles from the New York/Pennsylvania state line to the city of 

Sunbury. A TMDL for a portion of the upper Susquehanna River basin has been completed 

and identifies an average fish tissue value of 860 ppb. The upper limit for unrestricted fish 

consumption is 50 ppb (PADEP 1999). PADEP is actively monitoring the PCB levels in fish in the 

lower section from Sunbury to the Bay (2006, 2008, 2011, and 2012) and none of the results have 

resulted in a fish consumption advisory for PCBs. 

In Virginia, all five Bay tributaries have PCB fish tissue impairments with a total of 456 river 

miles and 2,011 square estuarine miles impacted. The James River has the highest number of 

impaired river miles (245 miles), whereas the mainstem Chesapeake Bay and its small coastal 

basins account for 79% of the impaired estuary footprint (VADEQ 2010). PCB TMDLs for the 

Potomac and Shenandoah Rivers have been completed (Haywood and Buchanan 2007; USEPA 

and VADEQ 2001). 

West Virginia identified the South Branch of the Potomac River as impaired in the 2010 report. 

In the Shenandoah Jefferson watershed, the Shenandoah is identified as impaired but with 

a completed TMDL (WVDEP 2010). As indicated in the above discussion, the TMDL for the 

Shenandoah was completed in collaboration with the Commonwealth of Virginia and the USEPA. 

The report did not identify any major West Virginia sources of the contamination but identified an 

industrial point source and a landfill as the major potential sources of PCB contamination to the 

Shenandoah River (USEPA and VADEQ 2001). In its January 2012 fish consumption advisory, the 

West Virginia Department of Health and Human Resources (WVDHHR 2012b) indicated that PCB 

levels in the Shenandoah may be declining. The average fish tissue concentration (skin-off) was 

250 ppb ww with a range from non-detect to 2,100 ppb (detection limit = 10 ppb). 
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2.0 EXTENT AND SEVERITY OF CONTAMINANT OCCURRENCE

All of the fish collected for skin-on analyses were below the WVDHHR screening value of 50 ppb 

ww (WVDHHR 2012a). The most recent Shenandoah data collected by the Commonwealth of 

Virginia (2005), however, indicates a continued need for fish consumption advisories because fish 

tissue concentrations continue to be elevated above the levels of concern (VADEQ 2012a). 

Within the entire Bay watershed, the Anacostia and Potomac Rivers in Washington, DC are the 

areas of greatest concern for PCB fish tissue contamination. The DDOE (http://ddoe.dc.gov/

service/fishing-district) currently advises the public not to consume any catfish, carp, or eels from 

waters of the District of Columbia due to PCBs and other chemicals. The most recent sampling 

in the Potomac and Anacostia Rivers within the District was conducted in 2007 (Pinkney 2009). 

Pinkney (2009) found that the highest total PCB concentrations were in American eel (Anguilla 

rostrata), where the median concentration was 2.18 ppm ww, over 100 times the USEPA (2000b) 

screening value of 0.020 ppm. One eel sample contained 4.00 ppm. Median concentrations in 

carp (Cyprinus carpio), channel catfish (Ictalurus punctatus), and blue catfish (I. furcatus) were all 

close to 0.80 ppm. 

Pinkney (2009) compared PCB fish tissue concentrations in 2007 with those measured in 2000 

(Pinkney et al. 2001a), using similar methods and fish with similar lengths. Median concentrations 

of PCBs in American eel, carp, and largemouth bass increased in both the Potomac and Anacostia 

Rivers whereas median PCB concentrations in channel catfish decreased in both rivers. Median 

PCB concentrations in sunfish decreased slightly. PCB concentrations were generally higher in 

the Anacostia vs. Potomac fish, but fish from both rivers were well above the screening limit 

thresholds.

Ecological concerns

Wenning et al. (2011) and Monosson (1999) reviewed literature on toxicological effects of PCBs 

on fish. Reported effects include mortality, impaired growth and reproduction, disruption of the 

endocrine and immune systems, biochemical changes, behavioral alteration, and mutagenicity. 

Iwanowicz et al. (2009a) documented adverse effects on the brown bullhead immune response, 

disease resistance and endocrine physiology following intraperitoneal exposure to 5 and 0.5 

mg/kg of the PCB mixture Aroclor 1248. Similarly, a significant negative correlation has been 

documented between PCB body burden and the immune response and endocrine physiology in 

wild-caught brown bullheads and largemouth bass (Iwanowicz et al. 2012). Barron et al. (2000) 

documented an increased prevalence of hepatic tumors and preneoplastic liver lesions in walleye 
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2.0 EXTENT AND SEVERITY OF CONTAMINANT OCCURRENCE

from the PCB-contaminated Green Bay area of Lake Michigan relative to a reference area. They 

stated that while these results did not show causation, they are consistent with studies that 

indicate that PCBs are liver tumor promoters in fish (Weisburger and Williams 1991). Wenning et 

al. (2011) concluded that data were inadequate to establish no-observable-effect concentrations 

(NOECs) based on PCB tissue residues.

Meador et al. (2002), in their review of salmonid toxicity data, suggested a total PCB tissue 

residue threshold of 2,400 ppb in lipid, which corresponds to 140 ppb ww tissue (D. MacDonald, 

MacDonald Environmental Services Ltd, personal communication). TAMS Consultants Inc. and 

Menzie-Cura Associates Inc. (2000) addressed the toxicological effects associated with PCB 

residues as part of the Hudson River Ecological Risk Assessment. Based on their literature 

review, they recommended a no-observable-adverse-effects level (NOAEL) of 1.9 ppm whole 

body weight and a lowest observable adverse effect level of 9.3 ppm body weight for total PCB 

concentrations. These NOAELs and LOAELs were applied to species that are resident to the 

Hudson River, most of which are also Chesapeake Bay species. Using a whole body to fillet ratio 

of 1.7 from Amrhein et al. (1999), these are converted to 1.1 ppm fillet (NOAEL) and 5.5 ppm fillet 

(LOAEL). Others have reported whole body: fillet ratios ranging from 1.7 to 3.1 (D. MacDonald, 

MacDonald Environmental Services Ltd, personal communication), which would lower the 

estimated fillet concentration thresholds.

Based on TAMS Consultants Inc. and Menzie-Cura Associates Inc. (2000) and Meador et al. 

(2002), it is reasonable to suggest that total PCB fillet concentrations above about 1.0 ppm may 

be associated with adverse biological effects in Bay watershed species. Such concentrations 

have been reported in bottom-dwelling fish in urban areas such as the District of Columbia 

(Pinkney 2009; Velinsky et al. 2011) and near USEPA National Priority List sites where PCBs are a 

contaminant of concern, such as the Marine Corps Base Quantico (Pinkney and McGowan 2006).

Since 1986, NOAA’s Mussel Watch program has monitored PCBs in shellfish within the Chesapeake 

Bay and its tributaries at five locations in Maryland and five locations in Virginia, with data 

summarized in Kimbrough et al. (2008). In Maryland, 2004-2005 data for oysters (Crassostrea 

virginica) at the five locations were included in the report with concentrations of 21, 23, 60, 64, 

and 79 ppb. In Virginia, 2004-2005 total PCB concentrations at the five locations were 21, 21, 50, 

57, and 157 ppb. Kimbrough et al. (2008) characterized concentrations as low, medium, or high 

as part of a nationwide comparison. The three highest Maryland concentrations, and the 50 and 
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2.0 EXTENT AND SEVERITY OF CONTAMINANT OCCURRENCE

57 ppb in Virginia, were characterized as medium concentrations. The 157 ppb concentration in 

Virginia was characterized as high. 

Wildlife

No recent (post 2000) reports of PCB residues or effects in mammals were identified. In general, 

birds are more tolerant to acute exposure to PCBs compounds than mammals, but a range of 

effects (e.g., enzyme induction, altered growth and reproduction, chick edema disease, immune 

dysfunction and endocrine disruption) have been linked to exposure (Rice et al. 2003). 

Threshold effect concentrations are based on residues in eggs or from blood samples and vary 

according to species. For total PCBs, estimated egg residue thresholds for impaired hatching or 

fledging success is 35,000 ppb ww in raptors and 23,000 to 142,000 ppb in terns, and thresholds 

for impaired 3- to 5- year productivity in raptors is 25,000 ppb (reviewed in Harris and Elliott 2011). 

The nestling eagle blood threshold for reproductive success is 189 ppb (Elliott and Harris 2002; 

Henny and Elliott 2007).

In the past 20 years, total PCBs concentrations have only been reported for a few bald eagle eggs. 

A single addled egg collected at Aberdeen Proving Ground in Harford County, Maryland in 2008 

contained 33,690 ppb ww (Mojica and Watts 2008), and two addled eggs collected from the Naval 

Support Facility Indian Head in Charles County Maryland in 2008 and 2009 contained 18,400 and 

18,300 ppb (Mojica and Watts 2011). Threshold effects for reduced productivity in bald eagle 

(Haliaeetus leucocephalus) have been estimated to be about 25,000 ppb ww of egg (Elliott and 

Harris 2002; Henny and Elliott 2007; Harris and Elliott 2011). Concentrations of total PCBs from 

blood samples of 58 nestling eagles from these same sites ranged from 7 to 106 ppb ww (Mojica 

and Watts, 2008, 2011), and are below the toxicity threshold for impaired reproduction of 189 ppb 

(Elliott and Harris 2002; Henny and Elliott 2007).

In a large-scale osprey study conducted in 2000 and 2001, total PCB concentrations in eggs 

collected from Baltimore Harbor and the Patapsco River, and the Anacostia and middle Potomac 

Rivers averaged 7,250 and 9,280 ppb ww, respectively (Rattner et al. 2004). The upper extreme 

value was 19,300 ppb from an egg collected near the Naval Research Laboratory on the middle 

Potomac, and was actually similar to the greatest historical values reported in osprey eggs from 

the Chesapeake (Wiemeyer et al. 1988). Osprey eggs from the Elizabeth River, the location of the 

largest naval port in the world, contained the lowest total PCB value, averaging 3,600 ppb. Total 
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2.0 EXTENT AND SEVERITY OF CONTAMINANT OCCURRENCE

PCB concentrations in eggs from the South, West and Rhode Rivers reference area averaged 4,600 

ppb, and ranged up to 12,400 ppb. Concentrations of 15 arylhydrocarbon (Ah) receptor-active PCB 

congeners (but not dioxins or dibenzofurans) were also quantified in these eggs. Concentrations 

of the toxicologically most potent coplanar and semi-coplanar congeners (i.e., congeners 77, 81, 

105, 126 and 169) did not differ much between study sites. Dioxin toxic equivalents (TEQs) of 

15 Ah receptor-active congeners did not differ among sites in this Chesapeake Bay study, with 

average site values ranging from 0.0545 to 0.218 ppb TEQ ww. Concentrations of total PCBs and 

dioxin TEQs in these osprey eggs were slightly above the no-observed-effect-level (0.136 ppb 

TEQs ww) for hatching success (Woodford et al. 1998; Harris and Elliott 2011).

Total PCBs were quantified in 22 addled peregrine falcon eggs collected between 1993 and 2002 

from locations in the Chesapeake Bay, and ranged from 3,460 to 12,500 ppb (Potter et al. 2009). 

It is difficult to assess the importance of these residues because PCB toxicity thresholds have not 

been rigorously developed for peregrine falcons. 

Findings of high concentrations of PCB congeners and toxic equivalents, as well as cytochrome 

P450 induction in Baltimore Harbor black crown night herons (Rattner et al. 1997), was the 

impetus for testing the hypothesis that PCBs might be leading to the declining size of the 

Baltimore Harbor heron colony (Rattner et al. 2001). Although total PCBs, 12 Ah receptor-active 

PCB congeners and dioxin toxic equivalents were up to 35 times greater in sample eggs from 

Baltimore Harbor compared to those from the reference area in the southern Chesapeake 

(Holland Island), overall nest success (0.74) and productivity (2.05 young/hen) were adequate to 

maintain a stable population. Furthermore, no significant relation was found between hatching, 

fledging and overall reproductive success and concentrations of PCBs and toxic equivalents. It 

was concluded that contaminants were not having a dramatic effect on reproduction in the 

Baltimore Harbor heronry.

In a preliminary study examining potential endocrine disruptive effects of PCBs, common tern 

eggs collected in 1994 from South Sand Point, off of Barren Island, contained relatively low 

concentrations of Aroclor 1260 concentrations ranging from 440 to 1,500 ppb ww (J.B. French, 

USGS, unpublished data). In testing this hypothesis, eggs were subsequently collected from 

Bodkin Island (Chesapeake Bay) which served as a comparative reference site for the more 

contaminated samples from Ram Island in Buzzards Bay, Massachusetts. Total PCBs concentrations 

(<10,000 ppb lipid weight) were much lower at Bodkin Island compared to Ram Island, but no 
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2.0 EXTENT AND SEVERITY OF CONTAMINANT OCCURRENCE

evidence was obtained to suggest that PCBs were evoking toxic effects in embryos 

(French et al. 2001). In 2010, common tern eggs were collected from Poplar Island for a flame 

retardant toxicity study, and six that were chemically analyzed were found to have very low 

total PCB concentrations (range: 310 to 440 ppb ww) and well below suspected adverse effect 

thresholds for reproductive success (Harris and Elliot 2011).

In portions of the Chesapeake, including USEPA-designated regions of concern (Baltimore Harbor, 

Elizabeth River, and Anacostia River) there is clearly evidence that total PCBs and Ah-receptor 

active PCB congeners induce cytochrome P450-associated monooxygenases, and may evoke 

adverse effects on some sensitive species of birds (Rattner et al. 1997; Rattner and McGowan 

2007). In a few circumstances, PCBs concentrations in eggs are great enough to contribute to the 

failure of eggs to hatch (Mojica and Watts 2008, 2011).

Conclusions

Existing data documenting the concentrations of PCBs in fish tissue in the Chesapeake Bay 

watershed are adequate to characterize the extent of PCB contamination in the watershed 

as widespread in sediment and fish tissue. Unlike concentrations of chlorinated pesticides, 

concentrations of PCBs in tissues of many species of Chesapeake Bay wildlife have not declined 

since the final USEPA rule restricting the manufacture, processing, and distribution of these 

compounds became effective in 1979. The environmental persistence of PCBs is an impediment 

to quick remediation of the impairments as legacy deposits remain a primary source of PCBs. 

However, continued authorized uses and the inadvertent production of PCBs remains a potential 

source of environmental exposure. Due to the widespread presence of fish consumption 

advisories within the Chesapeake Bay watershed, it is clear that PCB contamination is negatively 

impacting the use of watershed resources. 

2.2 Dioxins and Furans

Abstract

Dioxins and furans is the abbreviated name for a family of toxic substances that share a similar 

chemical structure, containing one to eight chlorine atoms attached to the carbon atoms of the 

parent chemical (dibenzodioxin and dibenzofuran). Dioxins and furans have no known technical 

use and are not intentionally produced. They are primarily released to the environment during 

combustion of fossil fuels (coal, oil, and natural gas) and wood, and during incineration processes 

(municipal and medical solid waste and hazardous waste incineration) and have been shown to 
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2.0 EXTENT AND SEVERITY OF CONTAMINANT OCCURRENCE

be present as contaminants in certain chlorinated pesticides (e.g., 2,4 -D). The few data available 

were collected from areas associated with a known potential for dioxin contamination. 

There are three sites within the Bay watershed with documented dioxin contamination in 

sediment: the Atlantic Wood superfund site located in the Elizabeth River, and two areas within 

Maryland’s North Branch of the Potomac River (maximum values of 2.674 ppb in the Elizabeth 

River and 0.2966 ppb in the North Branch of the Potomac). The latter two sites appear to be 

associated with historical dioxin contamination associated with a papermill in Luke, Maryland. 

Though data on dioxin contamination in the watershed are limited, available information 

indicates that the extent of dioxin contamination within the Chesapeake Bay watershed is 

localized. Impairments as a result of elevated levels of dioxin in crab and fish tissue have been 

identified in two areas of the watershed: the Elizabeth River and the C&D Canal. Whereas 

the Elizabeth River fish consumption advisory only limits the consumption of blue crab 

hepatopancreas, the C&D canal fish consumption advisory impacts the consumption of finfish. 

These localized fish consumption advisories impact the use of the Bay resources identified in the 

advisories. The severity of effects from dioxin is identified as localized. 

Background

Dioxins and furans is the abbreviated name for a family of persistent, bioaccumulative and 

toxic substances that all share a similar chemical structure. These chemicals contain one to 

eight chlorine atoms attached to the carbon atoms of the parent chemical (dibenzodioxin and 

dibenzofuran). The chlorinated dibenzo-p-dioxins (CDDs) include 75 individual compounds 

and the chlorinated dibenzofurans (CDFs) include 135 compounds. These individual 

compounds are referred to as congeners. The most widely studied of these compounds, 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD), is one of the most toxic to mammals and has 

received the most attention. Often referred to as “dioxin”, 2,3,7,8-TCDD serves as a reference 

compound for this class of compounds. The chemicals with properties similar to 2,3,7,8-TCDD 

are called “dioxin-like” compounds. Only seven of the 75 congeners of CDD and 10 of the 135 

congeners of CDFs are thought to exhibit “dioxin-like” toxicity (ATSDR 1994, 1998). 

Dioxins and furans have no known technical use and are not intentionally produced. They arise 

through inadvertent production in processes involving chlorine in uncontrolled reactions. Air 

emissions of dioxins and furans result from the combustion and incineration of fossil fuels, 

municipal and medical solid waste, and hazardous waste (ATSDR 1994, 1998). Additional 

sources include the burning of many materials that may contain chlorine, such as plastics, wood 
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2.0 EXTENT AND SEVERITY OF CONTAMINANT OCCURRENCE

treated with pentachlorophenol (PCP), pesticide-treated wastes and PCB-containing materials. 

They are also created in the pulp and paper industry from a process that bleaches the wood 

pulp (ATSDR 1998). Pesticides like 2,4-D have been shown to be contaminated with dioxins. 

Though manufacturing has been optimized to reduce the formation of impurities, the use of 

these pesticides still represents a potential source of dioxins (USEPA 2005). Other pesticides like 

triclosan, when in water, can also degrade to different forms of dioxin naturally in sunlight (Latch 

2000). USEPA (2001a) breaks down the sources into five types: combustion and incineration; 

metals smelting; refining and processing; chemical manufacturing and processing; reservoir 

sources, and biological and photochemical processes. Within the last few decades, regulations on 

incineration processes have resulted in a reduction of dioxin air emissions (USEPA 2001a). 

Dioxin exposure in waterways can be widespread, resulting from the dispersion of soil particles 

in erosion and run-off, volatilization from land and water, and resuspension of sediment particles. 

There is some uncertainty as to whether the atmospheric deposition represents contemporary 

inputs of dioxins or whether this deposition component is resulting from transfers among the 

environmental media (USEPA 2001a).

Dioxins and furans are found in the environment together with other structurally related 

chlorinated chemicals. Therefore, the Toxicity Equivalent Factor (TEF) methodology was 

developed to estimate the risk from exposure to these mixtures. Because very limited data on the 

toxicity of these chemicals exist, TEFs were developed and validated in studies in animals. TEFs 

are the result of expert scientific judgment using all of the available data and taking into account 

a variety of uncertainties. The TEF approach compares the relative toxicity of individual congeners 

to that of 2,3,7,8-TCDD, which is the most extensively studied. 2,3,7,8-TCDD was assigned a TEF of 

1.0 and the TEF for the other congeners range from 1.0 to 0.00001. The toxic potency of a mixture 

of congeners (i.e., the Toxicity Equivalents, or TEQ) is the sum of the products of the TEFs for each 

congener and its concentration in the mixture (ATSDR 1994, 1998).

Dioxins and furans can cause a number of health effects with complex mixtures identified as “likely 

human carcinogens” (USEPA 2001a). At ten times greater than background exposure, humans may 

experience changes in markers of early development and hormone levels. High dose exposures 

have caused chloracne skin disease. Animal studies show that animals exposed to dioxins and 

furans experienced changes in their hormone systems, changes in the development of the fetus, 

decreased ability to reproduce, and suppression of the immune system (USEPA 2012c).
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Water and Sediment

No water column data were found in the sources of information reviewed; this is not unexpected 

because dioxins and furans are hydrophobic and tend to sorb to soil and sediments. Additionally, 

dioxins may not be commonly assessed in jurisdiction waters. Water quality standards for human 

health protection exist in each of the jurisdictions and range from 5x10-9 to 5.1x10-7 μg/L. 

Dioxin/furan contamination was noted at three locations. One is in the Southern Branch of the 

Elizabeth River in Portsmouth, VA (approximately 12 miles from the Chesapeake Bay), the location 

of the Atlantic Wood Industries (AWI) Superfund site. One of the compounds used at the AWI 

facility as a wood preservative, pentachlorophenol, has been reported to contain dioxin and furan 

impurities. Sediment sampling conducted as part of the Superfund investigation showed that 

the dioxin in samples included predominantly octachloro dibenzo-p-dioxin (OCDD), with minor 

amounts of heptachloro dibenzo-p-dioxin (HpCDD). Sampling conducted in shallow sediment 

in the vicinity of the AWI site had a maximum TEQ of 2.674 ppb. Sediment sampling also showed 

detections of dibenzofurans, but at lower levels (USEPA 2007a). Remediation of the site includes a 

plan to dredge contaminated sediments. 

Another area with reported dioxin contamination is in the North Branch of the Potomac River. In 

the late 1980s and early 1990s, dioxin contamination in the North Branch of the Potomac River 

resulted from the discharge from the Westvaco paper mill in Luke, Maryland. Fish consumption 

advisories were issued at that time by both Maryland and West Virginia for a portion of the 

Potomac between Luke Maryland and Paw Paw, West Virginia (Wheeler 1992). Westvaco 

implemented control measures to reduce dioxin levels in the discharge and fish advisories were 

discontinued in the early 1990s. Given the hydrophobic nature of dioxins, the sediment may still 

be contaminated. However, no information was available on any sediment concentrations close 

to this discharge (Wheeler 1992).

The third area where dioxin contamination has been reported is also in the North Branch 

Potomac River. In Cumberland, Maryland, sediment sampling was conducted in 2009 as part 

of the Cumberland Dam Removal Feasibility Study (Princeton Hydro 2010). Sediment samples 

collected from within the dam and upstream detected dioxin-like compounds with the 2,3,7,8-

TCDD concentration in each sample ranging from 0.57 to 0.93 parts per trillion (ppt). The sum 

of dioxin-like compounds detected in sediment samples ranged from 99.5 to 296.6 ppt. There 

was no information regarding the source of contamination, however, the report notes that the 
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2.0 EXTENT AND SEVERITY OF CONTAMINANT OCCURRENCE

dam is approximately 28 miles downstream from the Westvaco paper mill in Luke, Maryland. The 

report recommended collection of additional information; however, no further studies have been 

conducted since that study was completed (Princeton Hydro 2010). 

To put these sediment values in context (maximum values of 2.674 ppb in the Elizabeth River and 

0.2966 ppb in the Potomac), the USEPA (1993) developed guidelines for evaluating potential risk 

to fish species. Low level risk is estimated at 0.060 ppb. Low level risk was based on a no effects 

threshold for reproductive endpoints. At 0.100 ppb, there is a high risk to sensitive species. These 

high risk concentrations are “derived from TCDD doses expected to cause 50 to 100% mortality in 

embryos and young of sensitive species”. 

 

Fish and Shellfish

In their water quality assessment reports, the states use fish advisories as the basis for listing 

streams as impaired. The fish advisories are coordinated with the respective states’ health 

departments. Only two states, Virginia and Delaware, listed streams in the Chesapeake Bay 

watershed, as impaired by dioxins. The listings are reported as “Dioxin”, or “Dioxin, including 

2,3,7,8- TCDD”. Specifically, Delaware listed the Chesapeake and Delaware Canal (C & D Canal), 

from the Maryland line to Delaware River (15 miles). This listing was based on an advisory for no 

consumption of all finfish due to dioxin, among other contaminants. The principal toxicological 

driver for this listing is to limit human health exposure to fish tissue contaminated with PCBs (R. 

Greene, DNREC, personal communication). No source for the dioxin is specified for this segment. 

VADEQ listed the Elizabeth River Southern Branch and its tidal tributaries (an area of 23 river 

miles) on the basis of a fish advisory issued for no consumption of blue crab hepatopancreas. 

The edible tissue, however, is identified as safe for consumption. As stated above, dioxin 

contamination was reported in the Southern Branch of the Elizabeth River as a result of 

contamination from the AWI Superfund Site. As part of the Superfund investigation, shellfish 

sampling was conducted as follows (results in parenthesis are maximum detections): 

» crab meat and whole crab (crab meat and hepatopancreas) collected adjacent to the AWI site 

(total dioxin TEQ 0.00026 ppb and 0.012 ppb, respectively); 

» oyster meat from oysters collected adjacent to the AWI site (0.0014 ppb); 

» crab meat and whole crab (crab meat and hepatopancreas) from near Scuffletown Creek 

(0.000023 ppb and 0.0042 ppb, respectively) (USEPA 2007a).
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2.0 EXTENT AND SEVERITY OF CONTAMINANT OCCURRENCE

As indicated above, two streams in the Bay were listed as impaired by dioxins, one in the Southern 

Branch of the Elizabeth River and one in the C & D Canal based on fish advisories. The fish 

advisories were issued due to the exceedance of each of the states’ benchmarks, the purpose of 

which is to minimize human health exposure through the consumption of fish tissue. The states’ 

schedule for issuing TMDL plans for these impairments is from 5 to 10 years. In the near-term, 

however, Superfund remediation is expected to address contamination in sediments near the 

AWI Superfund site in the Southern Branch of the Elizabeth River (USEPA 2007a). USEPA (1993) 

identified 50 ppt fish tissue concentration as a low risk threshold for effects in fish. For pisciverous 

mammals and avians, the low risk thresholds for fish tissue concentrations are 0.7 ppt and 6 ppt, 

respectively. High risk levels for effects in sensitive species were identified as 80 ppt for fish, 7 ppt 

for pisciverous mammals, and 60 ppt for pisciverous fish.

Wildlife

Apparently, only one study has examined the concentrations of dioxins and dibenzofurans in 

Chesapeake Bay wildlife. Addled and post-term peregrine falcon eggs were sampled for dioxin 

(2,3,7,8-TCDD: ranging from not detected to 97 ppt ww, not adjusted for moisture loss) and 

dibenzofurans (2,3,7,8-TCDF: ranging from not detected to 146 ppt ww, not adjusted for moisture 

loss) (USFWS et al. 2004). These data, in combination with coplanar PCB congener levels, were 

subsequently concentration-corrected for moisture loss to estimate their fresh weight values 

(Clark et al. 2009). These concentration values were used to estimate dioxin toxic equivalents. 

Geometric means of the peregrine egg samples from middle Chesapeake Bay (n=9) and upper 

Chesapeake Bay (n=2) were at or below the NOAEL of 0.23 ppb TEQ ww derived for the American 

kestrel (Falco sparverius) (USEPA 2003a). 

Conclusions

Existing data documenting the environmental occurrence of dioxins and furans in the 

Chesapeake Bay watershed are limited though sediment and fish tissue contamination has 

been identified near areas associated with historical contamination. Specifically, sediment 

contamination has been identified at the Atlantic Wood Superfund site located in the Elizabeth 

River and at two sites in the North Branch of the Potomac River downstream of a Luke, Maryland 

papermill. Dioxin (as measured by 2,3,7,8 TCDD) has been found to be elevated in fish tissue 

resulting in impairment listings in the Elizabeth River and the C&D Canal. An advisory restricting 

the consumption of blue crab hepatopancreas was issued for the segment of the Elizabeth River 

associated with the Superfund site (no consumption allowed). The C&D Canal also has a limit on 
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2.0 EXTENT AND SEVERITY OF CONTAMINANT OCCURRENCE

the quantity of finfish that can be consumed, though the principal driver for this advisory is to 

limit human exposure to PCB contamination. These listings have resulted in localized extent and 

severity of contamination. 

2.3 Polycyclic Aromatic Hydrocarbons

Abstract

Polycyclic aromatic hydrocarbons (PAHs) are a class of hundreds of chemicals composed solely 

of hydrogen and carbon in structures containing two or more benzene rings. In environmental 

surveys, scientists often refer to 16 USEPA priority PAHs (acenaphthene, acenaphthylene, 

anthracene, benz(a)anthracene, benzo(a)pyrene, benzo(b)fluoranthene, benzo(ghi)perylene, 

benzo(k)fluoranthene, chrysene, dibenz(a,h)anthracene, fluoranthene, fluorene, indeno(1,2,3-

c,d)pyrene, naphthalene, phenanthrene, and pyrene). Other researchers examine as many as 

34 compounds including these 16 and alkylated PAHs. PAHs are either petrogenic (derived 

from petroleum and coal) or pyrogenic (derived from the combustion of fossil fuels or wood 

products). PAHs are detected more often in sediments than in other media. Whereas PAHs 

can bioaccumulate in invertebrates such as mussels and clams, they are rapidly metabolized 

in fish and, therefore, are less frequently measured in fish-tissue monitoring programs than 

PCBs and metals. PAHs are widely detected in sediments in the Chesapeake Bay watershed at 

varying concentrations, with the highest reported in or near Baltimore Harbor, Anacostia River, 

and Elizabeth River. In the Elizabeth River, sediment remediation projects, with the aim of 

reducing contaminant concentrations (including PAHs) and associated biological effects, have 

been completed or are in progress. Liver tumors in bottom-dwelling fish from the Anacostia 

and Elizabeth Rivers have been statistically associated with exposure to PAH-contaminated 

sediments and are used to monitor conditions. Three jurisdictions (Maryland, Virginia, and DC) list 

impairments. The severity of effects due to PAHs is identified as localized. 

Background 

Polycyclic aromatic hydrocarbons (PAHs) are a class of hundreds of chemicals, composed solely 

of hydrogen and carbon in structures containing two or more benzene rings. In environmental 

surveys, scientists often refer to 16 USEPA priority PAHs (acenaphthene, acenaphthylene, 

anthracene, benz(a)anthracene, benzo(a)pyrene, benzo(b)fluoranthene, benzo(ghi)perylene, 

benzo(k)fluoranthene, chrysene, dibenz(a,h)anthracene, fluoranthene, fluorene, indeno(1,2,3-

c,d)pyrene, naphthalene, phenanthrene, and pyrene) (USEPA method 610) Other researchers 

examine as many as 46 separate PAH compounds and groupings including these 16 compounds 
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and additional alkylated PAHs (Hartwell and Hameedi 2007). PAHs in water will readily adsorb 

to sediments (USEPA 2008b; Van Metre et al. 2006). They may persist in the environment, and 

some have half-lives of up to 5 years (Greenfield et al. 2004; Oros et al. 2007). There are two 

types of PAHs: petrogenic PAHs are found naturally in petroleum and coal, whereas pyrogenic 

PAHs are formed during the burning of gasoline, coal tar, aluminum, fuel oil, and other fossil or 

modern biomasses. Petrogenic PAHs tend to have lower molecular weights (typically < 4 carbon 

rings), whereas pyrogenic PAHs tend to have higher molecular weights (≥4 carbon rings). PAHs 

with higher molecular weights tend to have lower solubility (ATSDR 1995, NIH 2005). With their 

decreased solubility, the higher molecular weight compounds tend to be found in sediment, 

whereas the lighter compounds may be found dissolved in water. Studies indicate that the load of 

PAHs in Bay tributaries may have a larger proportion of pyrogenic (e.g., combustion byproducts) 

compounds than the load in the mainstem (Hartwell and Hameedi 2007). 

Human health effects of PAHs were summarized in the Agency for Toxic Substances and Disease 

Registry (ATSDR 1995) toxicological profile. Many PAHs are classified as probable human 

carcinogens by the USEPA and the International Agency for Research on Cancer. 

Water

In Maryland, two segments of the Patuxent River are listed as impaired by PAHs resulting from an 

oil spill in 2000. The April 7, 2000, oil spill resulted from a break in a Pepco pipeline; two segments, 

Craney Creek and Buena Vista, have yet to meet the Phase I or Phase II cleanup status. PAH 

contamination as a result of this oil spill resulted in impairment of 0.33 estuarine acres (MDE 2010).

In the District of Columbia, there are impairment listings for PAHs in more than 12 stream 

segments as part of the District’s organics TMDL (DDOE 2010). A study titled "Sediment 

contamination studies of the Potomac and Anacostia Rivers around the District of Columbia" 

was completed by the Interstate Commission on the Potomac River Basin in 1992. The study 

included the tidal basin. Results of this study showed higher levels of PAHs at sampled outfalls 

and storm sewers to the tidal basin than in basin sediments. Results did not indicate a specific 

outfall as the source. The study indicated that the primary source of these hydrocarbons likely 

was much more diffused and probably related to vehicular traffic. 
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There are no reported impairments resulting from PAH contamination in water in the other states 

within the Bay watershed (DNREC 2010, WVDEP 2010, PADEP 2010, NYSDEC 2010). In a USGS 

study of organic wastewater compounds in Pennsylvania stream water samples collected from 

2007 to 2009, PAHs were occasionally detected in concentrations ranging from 4 to 42 μg/L (Reif 

et al. 2012). In Virginia, impairments are based on fish tissue (VADEQ and Virginia Department 

Conservation and Recreation (VADCR) 2010).

Sediment

Sediment triad studies typically use synoptically collected samples that are analyzed for sediment 

chemistry and benthic macroinvertebrate community structure, and tested for toxicity as an 

approach for evaluating habitat quality (Chapman 1990). The CBP funded two sediment triad 

studies, Pinkney et al. (2005) and Fulton et al. (2007), to focus on Bay tributaries identified by 

USEPA (1999) as having insufficient data to characterize the extent and magnitude of sediment 

contamination. Additional triad studies to characterize areas with data gaps were funded by 

VADEQ and conducted by Roberts et al. (2002, 2003, and 2004). Only the sediment PAH data from 

those studies are reported here.

Pinkney et al. (2005) collected samples from tidal sections of the Bohemia, Elk, Northeast, and 

Severn Rivers in Maryland. Total PAH concentrations ranged from 0.717 to 16.9 ppm, substantially 

less than the ERM of 44.792 ppm. 

Fulton et al. (2007) collected samples from a total of 60 stations in five areas of the Bay for which 

data previously had been insufficient: the Chester River, Nanticoke River, Pocomoke River, Lower 

Mobjack Bay (Poquosin and Back Rivers), and South and Rhode Rivers. Among these locations, 

the total PAH concentrations were highest at stations in the upper section of the South River. 

The average total PAH concentration in the South River ranged from 2.723 ppm in the lower 

section to 4.704 ppm (just above the ERL of 4.022 ppm) in the upper section. Average total PAH 

concentrations in the other tributaries did not exceed the ERL values: Chester River, 0.784 ppm; 

Nanticoke River, 1.036 ppm; Pocomoke River, 0.890 ppm; and Rhode River, 0.777 ppm. The 

average total PAH concentrations in Lower Mobjack Bay ranged from 0.167 ppm in the Back River 

to 0.309 ppm in the Poquosin River. 

Roberts et al. (2002) studied the tidal freshwater areas in the James and Appomattox Rivers in 

2001. In the nine sediment samples analyzed, total PAH concentrations ranged from 0.955 to 
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5.785 ppm, with the highest six samples exceeding the TEC of 1.610 ppm and none exceeding 

the PEC of 22.800 ppm. In 2002, Roberts et al. (2003) collected 12 samples from Lower Mobjack 

Bay portion of the Chesapeake which included the area south of the York River and included the 

Poquoson and Back Rivers. PAHs were only detected in five samples; the maximum total PAH 

concentration was 0.184 ppm, far below the ERL of 4.022 ppm. In 2003, Roberts et al. (2004) 

collected a total of 14 sediment samples from the Mattaponi and Pamunkey Rivers. PAHs were 

detected in only sample, with a total PAH concentration of 0.040 ppm, far below the TEC. 

Hartwell and Hameedi (2007) reported the results of the NOAA sediment triad study of the 

Chesapeake Bay conducted at 210 sites in 1998, 1999, and 2001. They found that total PAH 

concentrations varied from just over 4 ppb to 22.6 ppm. They stated that concentrations at only 

one mainstem and six tributary stations exceeded the ERL, and no concentrations exceeded the 

ERM. They stated that the highest concentrations in tributaries were in Baltimore Harbor, the 

James and Elizabeth Rivers, and the mouth of the Patuxent River.

The objective of the Hartwell and Hameedi (2007) study was to conduct a broad-scale survey of 

the Bay; therefore, it did not focus on highly contaminated areas. Thus, in the three Regions of 

Concern (Baltimore Harbor, Anacostia River, and the Elizabeth River), only eleven samples were 

collected: two from Baltimore Harbor and nine from the Elizabeth River. Velinsky and Ashley 

(2001) analyzed surface sediment collected at 114 locations in the Anacostia River, including the 

Washington Channel, for contaminants in 2000. Total PAH concentrations averaged 22.6 ppm, just 

below the 22.8-ppm freshwater PEC (MacDonald et al. 2000a). The maximum concentration was 

90.4 ppm. Samples from 52 of the 114 locations had concentrations above the PEC. Therefore, 

based on a comparison of concentrations of total PAHs with the PEC, concentrations of total 

PAHs appear to be frequently detected at concentrations associated with adverse effects on the 

benthic community. However, in a sediment triad study of 20 locations with total PAHs ranging 

from 1.005 to 57.907 ppm, McGee et al. (2009) reported only sublethal toxicity at the station with 

the highest total PAH concentration and no evidence of sediment toxicity at the other 19 stations. 

The most extensive survey of Baltimore Harbor sediment chemistry and toxicity was conducted in 

1996 (Ashley and Baker 1999). Total PAH concentrations ranging from 90 ppb to 46.2 ppm (which 

exceeds the ERM) were reported for 80 sampling sites, with several of the highest in Bear Creek 

near the Sparrows Point industrial facility. More recently, EA Engineering, Science and Technology, 

Inc. (EA 2009, 2011) collected sediment, water, and tissue PAH concentration data in the Coke 
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Point Offshore Area near Sparrows Point. The mean total PAH concentrations in the 18 sediment 

samples collected in the Coke Point Offshore Area was 569 ppm (range: 5.97 to 7,354 ppm) which 

greatly exceeds the ERM of 44.792 ppm and the average total PAH concentration in Baltimore 

Harbor channels of 3.24 ppm (EA 2009). In a risk assessment, EA (2011) concluded that there were 

ecological risks to aquatic organisms from exposure to PAHs in the Coke Point Offshore Area 

based on exceedance of the probable effect level (PEL) screening benchmarks of 1.442 ppm for 

low molecular weight PAHs and 6.676 ppm for high molecular weight PAHs (Buchmann 2008; 

MacDonald et al. 1996). A PEL is defined as an estimate of the concentration above which adverse 

effects on the benthic community frequently occur. 

The Elizabeth River Project (2008; a grass-roots citizens’ organization dedicated to the cleanup 

of the Elizabeth River) reported that concentrations of PAHs in some areas of the Elizabeth River 

were as much as 1,000 times the average concentration in the Chesapeake Bay. Concentrations of 

individual and total PAHs at 16 locations in the Elizabeth River sampled in 2007 ranged from 0.736 

to 383.2 ppm (Vogelbein et al. 2008). The highest concentration was found in the South Branch 

near the Atlantic Wood Superfund site. This site, operated from 1926 to 1992 as a wood-treating 

facility, released creosote, PCP, metals, and dioxins into the Elizabeth River. Hartwell and Hameedi 

(2007) reported lower total PAH concentrations at the nine sites sampled in the Elizabeth River 

in 2001, ranging from 1.678 to 24.617 ppm but their survey did not attempt to sample highly 

contaminated locations.

Fish and Shellfish

Fish tissue advisories and impairments

In Virginia, fish tissue impairments are listed for 1 square mile of estuary (Pohick Creek, due to 

benzo[k]fluoranthene), 74 acres of lakes, and 7 river miles (Gold Mine Creek in the James River 

due to benzo(a)pyrene, benzo[b]fluoranthene, and/or benzo[k]fluoranthene) (VADEQ and 

VADCR 2010). VADEQ lists PAH tissue data for fish and invertebrates within the Chesapeake Bay 

watershed for 2000, 2001, 2003, 2005, and 2008 on its Web site (http://www.deq.state.va.us/

Programs/Water/WaterQualityInformationTMDLs/WaterQualityMonitoring/FishTissueMonitoring/

FishTissueResults.aspx). VADEQ uses a screening value of 15 ppb (as a PEC (potency equivalent 

concentration)) for total PAHs based on the sum of the following seven carcinogenic compounds, 

weighted according to the following relative potency factors: benzo(a)pyrene (potency =1.0), 

benzo(a)anthracene (0.145), benzo(b)fluoranthene (0.167), benzo(k)fluoranthene (0.020), 

chrysene (0.0044), dibenzo(a,h)anthracene (1.11), and indeno(1,2,3-c,d)pyrene (0.055). 
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In all the sampling years, no concentration exceeded the PEC. The highest PEC concentration in 

fish sampled in these years was 5.521 ppb in a 2005 composite sample of mummichogs (Fundulus 

heteroclitus) collected from St. Julian Creek in the Elizabeth River watershed.

Fish tissue samples were analyzed for PAH concentrations as part of monitoring conducted by 

Washington, DC to update fish tissue advisories (Pinkney et al. 2001, Pinkney 2009). In 2007, 

samples of the following species were collected from the Potomac and Anacostia Rivers: 

sunfish, largemouth bass, carp, blue catfish, channel catfish, and American eel (Pinkney 2009). 

PAH concentrations were compared with the USEPA (2000) screening value of 0.00547 ppm, 

which is based on the cancer slope factor for benzo(a)pyrene, the most potent PAH carcinogen. 

Concentrations were also evaluated with respect to Toxic Equivalent Factors (TEFs) by using the 

values of Nisbet and LaGoy (1992), which relate the toxicities of various PAHs to that of benzo(a)

pyrene. For the TEF approach, the screening values developed by the DNREC and the Delaware 

Department of Health and Social Services (DHSS) were used (DNREC-DHSS 2005). 

In general, the two- and three-ring PAHs such as naphthalene and anthracene were detected at

higher concentrations than the five- and six-ring compounds such as benzo(a)pyrene (Pinkney 

2009). Concentrations of PAHs in all samples except one sunfish sample exceeded the screening 

value of 0.00547 ppm. The highest total PAH concentrations by far were detected in carp and 

ranged from 0.0706 ppm in the lower Potomac River sample to 0.384 ppm in the upper Potomac 

River sample. The median concentration (0.184 ppm) was about 34 times the screening level. 

Median total PAH concentrations in the other species ranged from 0.0079 ppm in the single 

smallmouth bass sample (from the lower Potomac River) to 0.0814 ppm in American eel. By using 

the TEF approach, however, no concentrations in any of the samples for any species exceeded 

the DNREC-DHSS (2005) screening value of 7 ppb. Median TEF concentrations ranged from 0.002 

ppb in smallmouth bass to 0.129 ppb in carp. The highest TEF, 0.217 ppb, was measured in a carp 

sample from the upper Potomac River.

The District of Columbia’s advisory is not linked to a specific chemical but states that the advisory 

is due to presence of “PCBs and other chemicals” (http://green.dc.gov/service/fishing-district). 

In view of the uncertainty regarding the risk from PAHs in fish tissue, it is reasonable to attribute 

most of the risk from fish consumption in District of Columbia waters to PCBs and chlorinated 

pesticides rather than PAHs (based on data in Pinkney 2009). 
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Ecological concerns

Logan (2007) reviewed the ecotoxicology of PAHs on fish. Although PAHs can be detected in fish 

muscle tissue, these compounds tend to be metabolized by vertebrates and eliminated through 

the liver. Elimination results in detoxification but also the generation of genotoxic metabolites 

(French et al. 1996). Therefore the most common biomarkers used to monitor PAH exposure in 

fish are bile PAH-like metabolites (Leadley et al. 1999) and DNA adducts (Reichert et al. 1998).

Chronic exposure of fish embryos to PAHs may result in death, deformities, or decreased 

growth. Narcosis is believed to be an important mechanism of toxicity and is attributable to 

low molecular weight volatile PAHs (French-McCay 2002). There is also concern about sediment 

toxicity to bottom-dwelling organisms, which can be associated with phototoxicity (Barron 2007). 

A diminished benthic biomass could affect the survival, growth, and reproduction of their fish 

predators. 

There are documented cases of liver tumors in bottom-dwelling fish that have been causally 

linked with exposure to PAH-contaminated sediments (Myers et al. 2003; Baumann and 

Harshbarger 1998). In the Chesapeake Bay watershed, the clearest linkages between liver tumors 

in fish and PAH exposure are in the Anacostia (Pinkney et al. 2009) and Elizabeth Rivers (Vogelbein 

and Unger 2006). There are also genetic changes associated with long-term exposure to PAHs. In 

a laboratory study with Elizabeth River sediments, Ownby et al. (2002) documented an increased 

tolerance to the acute toxic effects of creosote. The authors found that tolerance was hereditable, 

indicating that it resulted from natural selection.

Pinkney et al. (2001b, 2004) used biomarkers of exposure and response, tumor data, and sediment 

data to evaluate the association between PAHs and liver tumors in brown bullheads. In both 

studies, the concentrations of biliary PAH-like metabolites were examined as an indicator of PAH 

exposure. Pinkney et al. (2004) also measured concentrations of polycyclic aromatic compound 

(PAC)-DNA adducts, which are bulky molecules attached to the DNA that serve as an indicator 

of response to PAHs. A specific pattern (diagonal radioactive zone (DRZ)) in the radiographic 

determination of these adducts is indicative of PAC adducts. Using logistic regression, the authors 

reported a statistical association between liver tumors and biliary PAH-like metabolites. They also 

reported equally high concentrations of PAC-DNA adducts in 1- to 2-year-old fish, which already 

had a 10 to 17% liver tumor prevalence, as in the 3-year-old and older fish, which had a 50 to 

68% tumor prevalence. Therefore, the younger age classes were likely to have a high prevalence 
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as they reach age 3 or greater. The finding of high concentrations of PAC-DNA adducts with a 

strong DRZ signal, elevated bile PAH-like metabolites, and elevated sediment PAHs (15–31 ppm 

total PAHs within 1 kilometer of the fish collection sites) provide strong evidence linking the liver 

tumors with exposure to PAH-contaminated sediments. 

Monitoring has been conducted over the past 25 years to evaluate the status of sediment 

contamination in the Elizabeth River using mummichog liver pathology. Vogelbein et al. 

(1990) reported total PAH concentrations as high as 2,200 ppm in sediments from a creosotecontaminated site. They reported that 35% of mummichogs collected near this site exhibited 

hepatocellular neoplasms and 73.3% of them had foci of cellular alteration, which are 

hypothesized to be pre-neoplastic lesions. They also noted a variety of other pathologies, 

including elevated prevalence of exocrine pancreatic (Vogelbein and Fournie 1994; Fournie and 

Vogelbein 1994) and vascular neoplasms (W. Vogelbein, VIMS, personal communication). In 

contrast, mummichogs from two relatively uncontaminated study sites exhibited no proliferative 

liver lesions or other pathologies. 

Vogelbein et al. (1990) demonstrated a strong positive association between sediment PAH 

concentrations derived from creosote spills and development of proliferative liver lesions. As 

a result, the VADEQ and the Elizabeth River Project adopted the mummichog as a sentinel 

of chemical contamination. Since the late 1990s, this fish has been used in a long-term field 

monitoring program in the Elizabeth River (Elizabeth River Project 2008). Results of these studies 

indicate that liver histopathology in mummichogs is an effective bioindicator of sediment 

chemical contamination (Vogelbein et al. 1997, 1999, 2008; Vogelbein and Zwerner 1999; 

Vogelbein and Unger 2003). The greatest liver disease prevalence was found in mummichogs 

from the most heavily contaminated sites and the lowest disease prevalence occurred in fish from 

the least contaminated sites. 

Currently, mummichog liver histopathology and sediment PAH measurements are being used in 

association with ongoing sediment remediation efforts in the Elizabeth River, such as the Money 

Point Project started in 2009. Remediation efforts included nearshore sediment dredging and 

sand capping, and construction of a new onshore salt marsh. The constructed inshore salt marsh 

appears to have created a refuge for aquatic life from the sediment chemical contaminants found 

just offshore. Remediation reduced the bioavailability of sediment PAHs, which has already 

caused a decrease in liver lesion prevalence at the site (Vogelbein and Unger 2011). 
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Since 1986, NOAA’s Mussel Watch program has monitored PAHs in shellfish within the Chesapeake 

Bay and its tributaries at five locations in Maryland and five locations in Virginia with data 

summarized in Kimbrough et al. (2008). In Maryland, 2004-2005 data for oysters at three of 

the five locations were included in the report with concentrations of 266, 306, and 481 ppb. In 

Virginia, 2004-2005 total PAH concentrations at the five locations were 200, 207, 232, 307, and 

1583 ppb. Kimbrough et al. (2008) characterized concentrations as low, medium, or high as 

part of a nationwide comparison. The 1583 ppb concentration was characterized as a medium 

concentration. 

Wildlife

From 1986 to 1988, 35 muskrats (Ondatra zibethicus) were trapped on the Elizabeth and 

Nanesmond Rivers in Virginia. Twenty-two of 35 carcasses had detectable concentrations of PAHs 

(naphthalene, phenanthrene, anthracene, fluorene, pyrene, and chrysene), and the muskrat with 

the greatest concentration (phenanthrene) contained 0.15 ppm dry weight (Halbrook et al. 1993). 

To the best of the authors’ knowledge, PAH concentrations have not been reported in other 

wildlife species in the Chesapeake Bay.

Conclusions

PAHs are widely detected in sediments in the Chesapeake Bay watershed at varying 

concentrations, with the highest reported in or near Baltimore Harbor, Anacostia River, and 

Elizabeth River. Concentrations of PAHs have also been detected in the water column and in fish 

tissues, although, in general, fish tend to metabolize rather than accumulate these compounds. 

There are localized impairments in Virginia (fish tissue), Maryland (water), and DC (water). 

Concentrations of PAHs in sediments of the Elizabeth River and the Anacostia River have been 

linked with a high prevalence of liver tumors in bottom dwelling fish. Fish lesion prevalence in the 

Elizabeth River has been used as a tool to monitor the success of sediment remediation projects. 

2.4 Petroleum Hydrocarbons

Abstract

Petroleum hydrocarbons are a mixture of several hundred chemicals originating from crude oil. 

The chemical components of this group of compounds vary considerably in chemical properties, 

affecting their bioavailability, fate and effects. The compounds vary with regard to solubility in 

water, with the lower molecular weight aromatic hydrocarbons (benzene, toluene, ethylbenzene, 

xylene (BTEX)) exhibiting a relatively high degree of solubility relative to the higher molecular 
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weight compounds. For petroleum compounds, states typically evaluate compliance with a 

narrative standard (i.e., no visible sheen) or with a numeric standard, such as that for oil and 

grease compounds (which include petroleum and non-petroleum-based oils). Data on reported 

spills and/or observable sheens are readily available, though concentration-based data are 

limited. The data that are available indicate that petroleum contamination within the Chesapeake 

Bay watershed is predominantly localized to areas with heavy boating or shipping activity. Two 

of the partner jurisdictions have documented water column impairments for the surrogate 

parameter, oil and grease. In Washingrton D.C., there are approximately 1.7 miles of impairment 

in the Anacostia River and Hickey Run. Pennsylvania documents a total of five miles of river 

impairments; three miles of the Conococheague River are impaired. These impairments indicate 

localized severity of effects. 

Background

Petroleum hydrocarbons are a mixture of several hundred chemicals originating from crude oil. 

The chemical components of this group of compounds vary considerably in chemical properties 

(i.e., some are characterized by relatively high vapor pressures, whereas others exhibit low vapor 

pressures). Compounds with high vapor pressures, such as those commonly found in gasoline, 

volatilize readily, whereas those with low vapor pressures tend to remain in liquid form. The 

individual compounds vary with regard to solubility in water, with the lower molecular weight 

aromatic hydrocarbons (BTEX) exhibiting a relatively high degree of solubility relative to the 

higher molecular weight compounds. The individual petroleum compounds have widely ranging 

organic carbon-water partitioning coefficients, indicating that some compounds partition 

preferably to solids and sediment that may be present in the water column, whereas others 

remain in water. Those compounds remaining in the aqueous fraction are highly mobile and 

can be transported considerable distances from the point of entry into the environment. Of the 

BTEX compounds, benzene is classified as a mobile compound whereas toluene, ethylbenzene, 

and xylene are classified as intermediately mobile. Upon exposure of these compounds to the 

environment, microbial populations capable of utilizing petroleum hydrocarbons as a carbon 

source begin to biodegrade them, provided that nutrients are also readily available (Horel et al. 

2012). Horel et al. (2012) determined that addition of inorganic nutrients to the sandy beaches 

in the northern Gulf of Mexico both stimulated and enhanced the biodegradation of these 

compounds following the Deepwater Horizon accident in 2010.
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Petroleum hydrocarbon contamination can arise from a number of sources, given the common 

use of petroleum-derived products (e.g., gasoline, kerosene, asphalt). Combustion and accidental 

releases of these products into the environment are common and are associated with the use 

of any tool, vehicle, or equipment that requires their use (DDOE 2003). An analytical method 

exists to estimate quantities of total petroleum hydrocarbons when assessing environmental 

prevalence; however, the more commonly used methods to assess risk to aquatic life and human 

health involve the measurement of surrogates such as the BTEX compounds or oil and grease. 

Chesapeake Bay jurisdictions assessing their waters for aquatic life protection commonly refer 

to either a narrative standard (“no visible sheen”) or a numeric standard for oil and grease 

promulgated by the USEPA in 1986. A water standard for total petroleum hydrocarbons has 

not been developed; however, this measurement of oil and grease includes plant-based oils as 

well as those derived from petroleum products. The measurement of oil and grease, like that of 

total petroleum hydrocarbons, captures a wide array of organic compounds, each with distinct 

physical, chemical, and toxicological properties (USEPA 1986). 

The information on impairments was derived primarily from the jurisdiction water quality reports 

(DDOE 2010, PADEP 2010). Quantitative data were extracted from a completed TMDL report 

(DDOE 2003). 

Water

Only two jurisdictions cite oil and grease impairments within the Chesapeake Bay watershed: DC 

and Pennsylvania (DDOE 2010; PADEP 2010). The Distict of Columbia identifies approximately 1.7 

miles as impaired (Anacostia River and Hickey Run); all but 0.6 miles have a completed TMDL. In 

Pennsylvania, approximately five river miles in the watershed are identified as impaired for this 

contaminant, with three miles in the Conococheague River. These impairments will not impact 

the Bay and will be remediated as per state regulatory requirements. 

Water-column data for oil and grease concentrations were generated in support of the Anacostia/

Hickey Run TMDL study. These concentrations ranged from less than detection (less than 5 mg/L) 

to 116 mg/L. Of the 37 samples collected during this effort, concentrations in five were at or 

above the water-quality standard of 10 mg/L. However, although the most recently approved 

integrated assessment report from the DDOE (2010) indicated the presence of a visible sheen in 

Hickey Run, it noted that the suspected sources are currently being addressed. 
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Petroleum is released to the environment through spills both small and large, including chronic 

releases of crankcase oil from vehicles and discharges from boat engines. The National Response 

Center (NRC) staffed by U.S. Coast Guard personnel is to be contacted upon discharge of a 

“harmful quantity” of oil, defined as “any quantity that violates state water quality standards, 

causes a film or sheen on the water’s surface, or leaves sludge or emulsion beneath the surface" 

(USEPA 2006). Minor spills occur frequently within the Bay watershed and are reported to the 

NRC. For the 5-year period from 2007 to 2011, USFWS personnel received 2,734 spill-response 

notifications from the NRC for the Chesapeake Bay region. Most spills were very small and none 

involved wildlife. Though the majority of these spills could not be attributed to a known source, 

many were associated with boating or shipping activity. Virginia had the most reports. The largest 

number of oil and hazardous chemical spill incidents were reported in the Elizabeth River area (S. 

Lingenfelser, USFWS, personal communication). 

Because of the wide variety of compounds present in the general category of “petroleums,” 

completing a risk assessment for the group as a whole is challenging. USEPA (1976) promulgated 

a water-quality criterion for oil and grease to protect aquatic life from the harmful effects of 

petroleum and nonpetroleum oils. The potential effects that were identified include sublethal 

effects on cellular activity as well as physiological processes affecting feeding and reproduction. 

These sublethal effects resulting from chronic exposure could occur when petroleum 

concentrations reach levels of 10 μg/L. In many cases, the most toxic and long-lived components 

of petroleum in the water column are PAHs. The toxicology of PAHs is complex and is discussed in 

more detail in a previous section of this report. 

Both petroleum and non-petroleum oils can cause deleterious physical effects on aquatic 

organisms by coating gill surfaces; by increasing biochemical oxygen demand in the water body, 

potentially leading to fish kills; and by combining with surface debris to form tarballs that settle 

out of the water column and may smother benthic organisms. Though the physical effects of 

non-petroleum-based oils are the same as those of petroleum-based compounds, vegetablebased oils are non-toxic (USEPA 1976).

Sediment and Fish

Sediment and fish tissue residue data for petroleums were not available. See the section of this 

report on PAHs for more information.
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Wildlife

Since the early 1990s, two moderate-sized oil spills have resulted in wildlife kills in the Chesapeake 

Bay. On March 28, 1993, a 36-inch high-pressure pipeline running from the Gulf of Mexico to 

Maine ruptured and released more than 400,000 gallons of No. 2 heating oil into Sugarland Run 

Creek and the Potomac River, affecting a 9-mile stretch. Twenty-three oiled birds, including 

wood ducks (Aix sponsa), Canada geese (Branta canadensis), mallards (Anas platyrhynchos), and 

a kingfisher (Megaceryle alcyon), were observed, 18 of which succumbed despite rehabilitation 

efforts (Research Planning, Inc. 1993). Undoubtedly, more birds were affected by this spill. 

Another oil spill on April 7, 2000, involved a pipeline rupture that released 126,000 gallons of No. 

2 and No. 6 fuel oil at the Potomac Electric Power Company Chalk Point Facility near Aquasco, 

Maryland. The spill spread to Swanson Creek, a tributary of the Patuxent River; 55 dead birds 

(principally waterfowl, but also osprey, herons, gulls, and terns) were found, and 109 birds were 

collected for rehabilitation (12 of these died) (Cardona et al. 2001). This spill occurred near osprey 

nests and, although 10 adult ospreys were observed to be oiled, there was no evidence of effects 

on reproductive success (Michel et al. 2001). In addition, some visible evidence of oiling from this 

spill was found on 10 snapping turtles (Chelydra serpentine), 7 diamondback terrapins (Malaclemys 

terrapin), 1 Eastern box turtle (Terrapene carolina), 18 common mud turtles (Kinsternon 

subrubrum), 2 Eastern painted turtles (Chrysemys picta picta), 1 garter snake (Thamnophis sirtalis), 

and 8 Northern water snakes (Nerodia sipedon) (Tri-state Bird Rescue and Research Inc., historical 

records, Newark DE). 

Conclusions

Existing data on the occurrence of petroleum spills or observable sheens are readily available 

through the NRC. However, concentration-based data are much more limited and typically 

present in the form of the surrogate parameter “oil and grease”. Review of the NRC data 

indicates that the extent of petroleum contamination within the Chesapeake Bay watershed 

is predominantly localized to areas with heavy boating or shipping activity. Two areas of the 

watershed (i.e., District of Columbia and PA) have documented water column impairments for oil 

and grease, indicating localized severity for this contaminant group. 
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2.5 Pesticides

Abstract

The USEPA defines the term “pesticide” to include many kinds of ingredients in products, such as 

insect repellants, weed killers, disinfectants, and swimming pool chemicals, which are designed 

to prevent, destroy, repel, or reduce pests. Unlike most of the other toxic contaminant groups 

(other than some PAHs) discussed in this report, pesticides are intentionally applied to land 

and water surfaces throughout the Chesapeake Bay watershed. Occurrence data on some 

pesticides in the watershed are available and therefore do not limit general interpretations about 

their extent and severity. For example, the extent of occurrence of herbicides, and specifically 

atrazine and its degradates, is identified as widespread. Insecticides such as aldrin, chlordane, 

dieldrin, DDE/DDT, heptachlor epoxide, mirex, and their degradates occur in localized areas. 

These pesticides persist for long periods of time in areas of usage. Data and research gaps exist 

for other pesticides including some current-use insecticides and fungicides, and consequently 

the extent and severity remains uncertain and cannot be evaluated at this time. The severity 

of some pesticide effects as determined by conventional benchmarking approaches has been 

documented in individual integrated state assessment reports by the number of river miles that 

are considered to be impaired (i.e. aldrin, chlordane, dieldrin, DDE/DDT, heptachlor epoxide, 

and mirex). For other pesticides, data and understandings of the potential environmental 

behavior and effects are not as complete for the low level concentrations typically detected in the 

environment. In addition, recent laboratory and field research indicate some pesticides can cause 

adverse sublethal effects on receptor organisms at environmentally relevant concentrations. 

Although there are data gaps for some pesticides in the watershed, there are a large number of 

potential sources (e.g. agricultural, residential, etc.). When considered with the emerging literature 

on the sublethal effects of some pesticides (e.g. atrazine) and their degradates, there is potential 

for adverse ecological effects in the Chesapeake Bay watershed. 

Background

Generally chemical or biological agents, pesticides are used for a range of governmental, 

residential, industrial, and agricultural applications. This report summarizes chemical pesticides 

only. Some pesticides target only one or a few species, whereas others are broad-spectrum 

and target all similar species. Pesticides can be classified or grouped by (1) similarities in the 

target organism(s) (e.g., herbicide, insecticide, fungicide, etc.); (2) similarities in the molecular 

structure (e.g., organophosphates, organochlorines, carbamates, etc.); or (3) the mode of action 

(e.g., acetylcholinesterase inhibition, inhibition of sterol biosynthesis, growth regulation, etc.). 
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Application rates and procedures vary greatly depending on the targeted organism(s), mode 

of action, chemistry, and other factors such as whether the pesticide is no longer in use. For 

example, some pesticides are applied before plants emerge, whereas others are applied after 

they emerge. Some are applied directly to plant or animal tissue, whereas others are applied 

to the soil or other physical components of habitat (e.g., open-water surfaces). Application 

procedures as well as delivery modes such as irrigation waters, aerial spraying, and tractor-applied 

dictate to a large extent how and where the active ingredients will be mobilized, deposited, 

and/or transformed. Changes in amounts of pesticides manufactured and used over time also 

influence their environmental occurrence and distribution. These pathways then determine the 

extent to which receptors (included target and nontarget organisms) are exposed. In order to 

facilitate application and uptake, various adjuvants are commonly used in pesticide mixtures. 

Consideration of pesticide adjuvants, although they are potentially important toxic chemicals, 

is beyond the scope of this chapter, which focuses on active pesticide ingredients and their 

degradation products. In addition, the USEPA definition of pesticide includes disinfectants. 

This broadens the scope of pesticides as a group of toxic contaminants and includes potential 

sources such as disinfectants and disinfection by-products in wastewater effluents and other 

environmental pathways. Those pesticides were beyond the scope of this section but are 

discussed elsewhere in the report (e.g. Household and Personal Care Products).

Agricultural uses account for about 75% of all pesticides, but at least one pesticide is stored 

in 85% of U.S. households, and one to five pesticides are stored in 63% (Delaplane, 1996), and 

pesticides are also used to control forest pests and for other purposes as well. Kiely et al. (2004) 

estimate a per capita home, garden, and personal-care pesticide loading of 0.42 pounds per 

year in the Chesapeake Bay watershed, which results in a loading to the watershed of about 6.5 

million pounds annually, not including agricultural uses. Hively et al. (2011) showed that the fate 

of herbicides (atrazine and metolachlor) is related to landscape features and characteristics in 15 

subwatersheds in the Choptank River system. Although agricultural and residential uses account 

for much of the pesticide use in the watershed, other land activities associated with golf courses, 

mosquito control, gypsy-moth control, control of invasive or non-native species, and weed control 

on rights-of-way are also potential sources of these contaminants. Therefore, virtually all land 

uses in the Chesapeake Bay watershed are potential sources of pesticides, and many provide 

direct pathways to the environment. 
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In 2009, a study (Maryland Pesticide Network 2009) of pesticides in the Chesapeake Bay 

watershed was published with input from a range of stakeholders and technical experts, 

including representatives of Federal and State agencies, and academia. The report noted that, 

although more data are needed, some pesticides have potential adverse effects on the watershed. 

Many previous studies have prioritized pesticides on the basis of combinations of occurrence 

(extent) and toxicity (severity) as potential environmental contaminants. For example, in 2006 the 

CBP Toxics Subcommittee assessed organic chemicals, including a range of pesticides previously 

documented to occur in the Bay, to occur in its tidal tributaries, or to have been used in the 

watershed with the potential to reach the Bay. The report further prioritized 35 contaminants, 

including 10 organochlorine and 2 organophosphate pesticides, for additional study on the basis 

of the following criteria: (1) the chemical appeared on lists of priority chemicals submitted by 

the Chesapeake Bay Program Toxics Subcommittee, (2) loading estimates for the chemical were 

available in the 1999 Chesapeake Bay Toxics Loading and Release Inventory (CBP 1999), and (3) 

detections of the chemical were reported in the 1999 Toxics Characterization Report (USEPA 

1999). These results represent only one attempt to prioritize pesticides in the watershed, with the 

acknowledgment that other prioritization schemes may yield different results. That prioritization 

is another indication that multiple pesticides warrant attention as potential environmental 

contaminants in Chesapeake Bay watershed.

There are many point and nonpoint sources of the pesticides that enter the Chesapeake Bay 

watershed. As previously mentioned, these chemicals are deliberately applied to the land and 

water surfaces. In addition to intentional applications, spills and other sources and indirect 

pathways to the environment, such as wastewater effluents (solid and/or liquid), landfill leachates, 

spray drift, runoff to surface water, and leaching to groundwater, provide inputs. These sources 

are all closely related to the usage and use patterns of pesticides. A multitude of pesticides, each 

with its own “footprint” in terms of geographic extent and location, is used in the Chesapeake 

Bay watershed. Pesticide usage changes over time as new products are registered, old ones are 

phased out, and pest-control needs and strategies evolve. There are many reasons for phasing 

out or modifying older pesticides in favor of new ones: loss of effectiveness due to increased 

resistance in target organisms; technological advances that enable the same level of effectiveness 

to be achieved with smaller quantities of pesticides (e.g., increases in potency of active 

ingredients); and changes associated with re-registration processes. For example, Hartwell (2011) 

notes that the total mass of pesticides being applied to the Chesapeake Bay watershed declined 

during the period 1985 to 2004; however, because the potency of these chemicals increased 
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during the same period, the “toxic units” (a measure of the toxicity of the active ingredients to 

a range of animal and plant species) remained approximately static or increased depending on 

the bioassay test organism(s) used. As a result of increases in usage and new analytical methods 

synthetic pyrethroids, are detected in increasing frequency in both surface water/suspended 

sediment and stream bed sediment (Kuivila et al. 2012). In addition, recent agricultural uses of 

pesticides have begun to shift toward the use of transgenic crops. Two of the most common 

examples include those that are resistant to targeted pesticides (e.g., Roundup Ready® 

soybeans) and/or are capable of creating pesticides through otherwise normal metabolic 

activities (e.g., “Bt corn”). In the Chesapeake Bay watershed Roundup Ready® soybeans are 

planted more extensively than corn with atrazine and metolachlor being used as pre-emergent 

herbicides on corn. As a result of transgenic crops, pesticide usage and geographic patterns 

are changing. For example, glyphosate herbicide usage increases where Roundup Ready® 

soybeans are grown, whereas usage of other herbicides (e.g., metolachlor) decreases. Toxins 

produced in situ by transgenic crops such as Bt corn are released to the environment in 

unquantified amounts, but recent studies have documented their occurrence and potential 

effects on aquatic invertebrates (Rosi-Marshall et al. 2007).

The fate and transport properties of pesticide formulations in the environment vary with active 

ingredient, adjuvant characteristics and purposes, application procedures, seasonality, and other 

factors (Barbash and Resek 1996). Generally, pesticides are biologically active agents capable 

of persisting long enough to be effective in environmental applications. Therefore, the halflife of all currently registered active pesticide ingredients ranges from hours to days to weeks 

to months depending on the nature of the molecule. Some pesticides are water soluble and 

can be expected to be mobile in aquatic environments, whereas others are hydrophobic and 

are likely to sorb to solids. Foster et al. (2000) documented that measured concentrations of 

organophosphorus pesticides in tributaries to the Chesapeake Bay varied mainly in response to 

the timing of their application, whereas the particle-reactive organochlorine insecticides were 

more directly correlated with stream discharge. Finally, although not a focus of the summaries 

in this report, atmospheric transport of pesticides has been shown by a number of studies across 

the globe to be an important ecological pathway for exposure (Majewski and Capel 1995). Kuang 

et al. (2003) detected metolachlor, atrazine, simazine, endosulfan, and chlorpyrifos in air and rain, 

with maximum concentrations during the time when crops were planted; they suggested that 

an atmospheric source from outside the watershed was likely contributing to some of the high 

values. Summarizing the environmental behavior of all pesticides is beyond the scope of this 
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report; however, information on half-life/persistence and the potential for the active ingredient 

to enter and move in aquatic environments are part of the registration process considerations 

required by USEPA. Many documented cases show long-term persistence of pesticides and/or 

their degradates (e.g., DDT/DDE; ATSDR 2002) as well as mobility in the environment (Barbash 

and Resek 1996). In addition, the environmental behaviors exhibited by the environmental 

transformation products of pesticides can be very different from those of the parent active 

ingredient (Barbash and Resek 1996).

Pesticide occurrence in ambient water resources, as residues on produce and other food products, 

and in our drinking waters has been documented by a number of nationally relevant studies. 

The USGS National Water-Quality Assessment (NAWQA) program has monitored and assessed 

the extent of pesticide occurrence in the Nation’s water resources since the early 1990s. These 

assessments have targeted a range of insecticides, herbicides, and some fungicides, and have 

shown close relations between pesticide applications and occurrence in nearby water resources 

throughout a range of hydrogeologic environments (Gilliom et al. 2006) Other factors such 

as the physiochemical characteristics of the pesticide molecules, the nature and properties of 

associated soils, and other hydrogeologic controls have been shown to be related to pesticide 

occurrence in water resources. Gilliom et al. (2006) note that the major findings of the NAWQA 

program with respect to pesticides include: (1) pesticides are frequently present in streams and 

groundwater; (2) pesticides are seldom present at concentrations that exceed human-health 

benchmarks; however, current benchmarks have been, and will be continuously, updated and 

revised as understandings of pesticide modes of actions in nontarget organisms, mixture effects, 

and other nonconventional toxicological effects are advanced; and (3) pesticides occur in many 

streams at concentrations that may have effects on aquatic life or fish-eating wildlife. The U.S. 

Department of Agriculture (USDA) publishes annual summaries of the Pesticide Data Program 

that include occurrence data for pesticide residues on a range of food commodities and drinkingwater sources in several states across the Nation, including some samples collected in Maryland, 

New York, Pennsylvania, and Virginia (USDA 2009). Although individual State data were not 

summarized for this report, nationally representative findings include: (1) 17% of food samples 

analyzed contained residue of only one pesticide; (2) 40% of food samples analyzed contained 

more than one pesticide; (3) 29 different pesticide residues (including metabolites) representing 

19 parent compound pesticides were detected in the 278 groundwater samples collected from 

farm wells, schools, and domestic wells; (4) 53 different pesticide residues (including metabolites) 

representing 42 pesticides were detected in the 306 finished municipal drinking-water samples 
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analyzed; and (5) 49 different pesticide residues (including metabolites) representing 38 

pesticides were detected in 306 samples of untreated (at intake locations) municipal drinking 

water. The detections of some pesticides cited above in sources of drinking water indicate the 

likelihood of at least one other potential exposure route that is unrelated to the workplace or 

application of pesticides. The presence of these contaminants in human tissue and drinking 

water does not necessarily indicate adverse health effects; however, in referring to low-level 

environmental exposures of pesticide residues from food sources (another exposure pathway 

for humans and animals), the National Institutes of Health notes that “scientists do not yet have a 

clear understanding of the health effects of these pesticide residues”. However, Birnbaum (2012) 

cited a growing body of research that links pesticide exposures to both acute and long-term 

health impacts including certain cancers, neurological, developmental, psychiatric, learning, 

behavioral and respiratory impacts, birth defects and immune system disorders.

As mentioned above, pesticides are a unique group of toxic contaminants because they are 

applied directly to land surfaces in a variety of land-use settings across the watershed. As a result 

of advances in modern detection technologies such as mass spectroscopy and gas or liquid 

chromatography and the range in their hydrophilic and lipophilic properties, pesticide residues 

(including parent and degradate compounds) are likely to be found in detectable concentrations 

in various environmental matrices including water, sediment, air, and fish tissue. Because the 

USEPA pesticide registration process requires testing and data, many sources of information 

about individual pesticide toxicities and potential behavior in the environment are available, 

particularly for the concentrations likely to be found as a result of pesticide applications in 

the field. This and other information sources (e.g., the Safe Drinking Water Act) have produced 

benchmarks to interpret environmental pesticide data. For example, the USEPA Office of 

Pesticides Program Aquatic Life benchmark for atrazine is 65 μg/L; however, Tillitt et al. (2010) 

showed that chronic exposures to concentrations as low as 0.5 μg/L will cause fathead minnows 

to spawn less frequently and to produce fewer eggs than controls. Recently, the usefulness of 

conventional “benchmarks,” which have been, and continue to be, developed on a contaminantby-contaminant basis for a limited range of toxicological endpoints, has been questioned 

(Feingold et al. 2010). In other recent research on the potential for pesticide usage to do 

environmental harm, the adjuvants have been implicated in potentially increasing the toxicity of 

pesticide mixtures in the environment (Brausch et al. 2007; Katagi 2008). 
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Impairments cited in the following paragraph are focused on pesticides that are no longer in 

use in the watershed. In Virginia, along the Potomac and Shenandoah Rivers, 7 river miles are 

impaired as a result of the insecticides heptachlor epoxide (five impaired segments) and chlordane 

in the James River Basin. Impairments are listed from chlordane (3 impaired miles) and from 

other insecticides (aldrin and DDE/DDT), totaling 9 impaired river miles; mirex, also an insecticide, 

impairs 5 river miles in the James River and 49 miles in the Rappahannock River (VADEQ and 

VADCR 2010). Two river segments in Maryland, the Back River and the Baltimore Harbor, and one 

impoundment, Lake Roland, are listed as impaired waters for chlordane. One location in the 

northwest portion of the Anacostia River in Maryland is listed as impaired by heptachlor epoxide 

contamination (MDE 2010). In the Potomac and Anacostia watersheds in the District of Columbia, 

chlordane, DDE, DDD, DDT, dieldrin, and heptachlor epoxide all contribute to impairments (DDOE 

2010). Pesticides are also responsible for the impairment of 17 river miles in the Susquehanna 

River Basin of Pennsylvania (PADEP 2010). New York and West Virginia do not report any pesticide 

contamination entering the Chesapeake Bay from their portions of the watershed (NYSDEC 2010, 

WVDEP 2010). 

Water

Many sources of data and information on pesticide extent in waters of the United States at local, 

regional, and national levels are available. Among the Federally-funded national programs that 

address pesticide occurrence in the environment, the U.S. Geological Survey’s NAWQA Program 

(Ator et al. 2005; Ator 2008; Ator and Denver 2012; DeBrewer et al. 2007, 2008; Denver et al. 2004, 

2010; Domagalski et al. 2008; Phillips et al. 2007) has found:

» Pesticides are widely detectable in non-tidal streams and groundwater of the Bay watershed in 

areas of use. 

» Pesticides are detectable year-round in many streams. Concentrations rarely exceed 1μg/L but 

can increase for example during, or shortly after, application if runoff transports contaminants 

to receiving water bodies.

» Concentrations of degradates of metolachlor, alachlor, and acetochlor often exceed those of 

parent compounds in groundwater and streams, in some cases by orders of magnitude. 

» Many pesticides (such as atrazine, metolachlor, and degradates) occur widely in groundwater, 

and groundwater serves as an important pathway for transport from the land surface to 

streams.
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» The extent of pesticide occurrence varies in time as new pesticide products come into 

use and others are no longer used. For example, concentrations of diazinon decreased 

significantly in Accotink Creek (a small stream in suburban Distict of Columbia) following 

usage restrictions in the early 2000s. The introduction of genetically modified crops and their 

associated pesticides into the Chesapeake Bay watershed has also changed pesticide usage 

over time. 

Water samples collected in Chesapeake Bay tidal waters during spring 2000 and late summer 

2004 were analyzed for a range of pesticides (McConnell et al. 2007). Of all pesticide groups 

analyzed for, agricultural herbicides were detected most frequently. Among the herbicides 

analyzed for, the parent compounds atrazine, simazine, and metolachlor were detected at 

maximum concentrations of 1.29, 0.49, and 0.61 μg/L, respectively. The highest concentrations 

of these compounds were observed in the Nanticoke, Pocomoke, and Chester Rivers on the 

Delmarva Peninsula, reflecting the intensity of agricultural production in the watersheds. 

Degradates of these parent compounds were also frequently detected, with the metolachlor 

degradation product, metolachlor ethane sulfonic acid (MESA), found in greater concentrations 

(2.9 μg/L) than the parent compound in almost all of the tidal regions assessed. Atrazine was 

detected in 100% of the samples collected in 2004. McConnell et al. (2007) noted that herbicides 

and associated products are present within both agricultural and nonagricultural tidal areas of 

the Chesapeake Bay throughout the year, even though they are used primarily in the spring. In a 

more detailed study of pesticides in upland areas of a Delmarva Peninsula agricultural watershed, 

outlets of 15 non-tidal sub-watersheds of the Choptank River were monitored approximately 

monthly under baseflow conditions from 2005 to 2007 (Hively et al. 2011). This work was part of 

a national experiment called the USDA Conservation Effectiveness Assessment Project (CEAP). 

Across all stations, the maximum atrazine, metolachlor and MESA concentrations were 7.6, 4.4, 

and 8.6 μg/L respectively with the greatest concentrations occurring between approximately 

April 1 and June 1 corresponding to corn planting activities. At most sites, only a moderate 

increase in metolachlor occurred after soybean planting in June, indicating that most soybeans 

planted were “Roundup-ready®” and did not include the use of metolachlor as a pre-emergent 

herbicide. Mean concentrations for the entire sampling period were 0.19, 0.29, and 4.2 μg/L for 

atrazine, metolachlor and MESA, respectively. 

Recognizing the need to develop estuarine criteria for low levels of pesticides in the Chesapeake 

Bay, Hall et al. (1997) studied the chronic toxicity of atrazine to Sago pondweed (Potamogeton 
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pectinatus) under varying conditions of salinity. This study showed that toxicity as well as 

bioavailability was influenced by salinity and highlights the typically subtle and complicating 

factors involved in determining conventional “benchmarks”. Since then, many other studies have 

concluded that conventional benchmarking approaches to decision making are inadequate for a 

variety of reasons. For example, Fatima et al. (2007) exposed goldfish to environmentally relevant 

mixtures of herbicides (atrazine, simazine, diuron, and isoproturon) and observed biomarkers 

indicative of immune suppression. They concluded that these common environmental 

exposures cause immune suppression in goldfish, thereby representing an endpoint and 

exposure scenario not quantified in conventional benchmarking studies. 

Whitall et al. (2010) analyzed water samples collected from seven sites in the Choptank River 

estuary on the Delmarva Peninsula during base flow and documented the seasonal signal in 

herbicide (atrazine, simazine, and metolachlor) concentrations known to dominate pesticide 

occurrence in streams throughout the Nation. Observed concentrations of individual herbicides 

did not approach established levels of concern (10-20 μg/L) for aquatic organisms (USEPA 

2006b) and were below the USEPA drinking-water standard for atrazine (3 μg/L). Total triazine 

concentrations (atrazine, simazine, CIAT, and CEAT), however, exceeded 3 μg/L in eight samples, 

highlighting the importance of assessing the effects of mixtures. Additionally, the degradation 

products (CIAT, CEAT, MESA, and MOA) were associated with groundwater pathways and were 

frequently found in higher concentrations than the parent compounds (Whitall et al. 2010). 

Hladik et al. (2005) also documented that degradates of triazine herbicides were found in higher 

concentrations than parent products in the upper Chesapeake Bay, and noted the importance of 

assessing degradate compounds in ecological studies.

Hall et al. (2009) studied the potential effects of irgarol (an algicide used as an anti-fouling agent 

in paints for boat hulls) and its major metabolite on phytoplankton in the Port Annapolis marina 

and Severn River system, Maryland. These sites were chosen to represent a range in use of irgarol 

including a high-use area (marina) and a background site (Severn River near its confluence with 

the Chesapeake Bay). In the study, a probabilistic approach was used to determine ecological risk 

by comparing exposure data with toxicity endpoints (NOEC). Field data could not confirm the 

microcosm NOEC in the high-exposure area (marina). 

As noted above, there are many local environmental studies that assessed pesticide occurrence in 

the Chesapeake Bay watershed. Recently, passive-sampler technologies have been used to detect 
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a broader range of pesticides in the Potomac River Basin (Alvarez et al. 2009). In general, pesticide 

occurrence data from several more recent studies including Ator and Denver (2012), Chambers 

and Leiker (2006), McConnell et al. (2007), Battaglin et al. (2009), Loper et al. (2007), and Hively et 

al. (2011) reflect changes in pesticide trends (e.g., increases in glyphosate use) as well as changes 

in sampling and analytical methods to include a wider range of degradate compounds. 

Kingsbury et al. (2008) collected raw and treated drinking water from nine community water 

systems ranging in size from a system serving about 3,000 people to one that serves about 2 

million people. The study included one site on the Potomac River where simazine, atrazine, 

DEA, and metolachlor were consistently detected throughout the year (i.e., 100% detection 

frequency). Other pesticides frequently detected include 2,4-D (58%), 2-hydroxyatrazine 

(92%), acetochlor (17%), benomyl (8%), carbaryl (33%), diazinon (33%), deisopropylatrazine 

(8%), 3,4-dichloroanaline (17%), diuron (25%), fipronil (17%), prometon (92%), MCPA (33%), and 

tryclopyr (17%). 

Sediment

The NOAA Contaminated Sediment report (Hartwell and Hameedi 2007) identifies 

organophosphate-type pesticides as being among the most common of all synthetic organic 

chemicals present in tidal sediment, with the largest concentrations reported for the Choptank 

and Nanticoke River watersheds in Maryland during spring runoff. Elevated levels of insecticides 

including chlordanes, heptachlors, nonachlors, aldrin, dieldrin, endrin, and endosulfan were 

found in sediments in the Elizabeth River in Virginia, and trace amounts of insecticides were found 

throughout the other tidal portions of the Bay. Hexachlorocyclohexane (HCH) concentrations 

were elevated in the Patuxent, Potomac, and Eastern Shore tributaries of Maryland. DDT was 

found throughout the Bay but was concentrated mainly in the upper Bay, and exceeded the 

ERM only in the Elizabeth River. Other pesticides detected were mirex, chlorpyrifos, butyltins, 

tributyltin (TBT), and chlorinated benzenes (Hartwell and Hameedi 2007). 

Organochlorine pesticides that are no longer in use in the watershed can be extremely persistent 

in soils and runoff events may move these soils into waterways where they can be accumulated 

in sediments and in wildlife. Areas that have historically been used for agriculture and residential 

areas that received high concentration applications of chlordane for termite control may be 

important sources of these compounds. In a recent study by Goel et al. (2010) concentrations of 

these “legacy” compounds were measured in air and rain from three different locations on the 
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Delmarva Peninsula over a four-year period. Overall, concentrations of these compounds were 

found to be slowly decreasing over time reflecting a gradual decrease in soil concentrations. 

However, results indicate significant local sources of the insecticide dieldrin and the DDT 

degradation product DDE in soils of the Delmarva Peninsula. Little is known regarding the 

location and magnitude of contamination from these historical sites which are likely distributed 

throughout the Chesapeake Bay watershed. 

Fish

Analyses of fish tissue for pesticide residues have been conducted for a variety of monitoring 

and research objectives throughout the watershed and it is beyond the scope of this report 

to summarize all these results. For illustrative purposes however it is instructive to highlight 

some selected results in order to provide further perspective on the extent and(or) severity of 

some pesticides in the watershed. For example, surveys in the District of Columbia (Pinkney 

et al. 2001a; Pinkney 2009) have shown elevated levels of chlordane in the tissue of certain fish 

species across many portions of the Anacostia and Potomac Rivers. Blazer et al. (2010) analyzed 

organ-specific fish tissue from fish in the Potomac River watershed; the following pesticides 

were measured in at least one tissue analyzed: chlorpyrifos, the phenyl pyrazole pesticides 

(fipronil, finpronilsulfide, and desulfynylfipronil), hexachlorobenzene, pentachloronitrobenzene, 

pentachloroanisole, oxychlordane, trans-chlordane, cis-chlordane, trans-nonachlor, cis-nonachlor, 

dieldrin, and DDx (DDE, DDD, DDT). One or more organochlorine pesticides were found in all 

tissues of both sexes. In addition, the ovary from a female smallmouth bass collected in the 

South Branch Potomac River contained 30.5 ppb total organochlorine pesticides, with 13.6 

ppb DDx; skin from the same individual contained 6.4 ppb total organochlorine, with 2.6 ppb 

from DDx. Pesticide compounds detected in individual organ tissue varied in presence and 

magnitude; muscle tissue generally contained the lowest concentrations. 

Ripley et al. (2008) discuss possible links between atrazine exposures and immune suppression 

in fish of the Potomac River system. A recent summary article by Rohr and McCoy (2011) showed 

potential behavioral and immune suppression effects when fish are exposed to environmentally 

relevant concentrations of pesticides and pharmaceuticals. In more recent work within the 

Chesapeake Bay watershed, Kolpin et al. (2013) have shown statistically significant correlations 

between atrazine concentrations in the water column above smallmouth bass nesting sites in 

the Potomac River and intersex conditions in male smallmouth bass collected at those sites (see 

Chapter 3.0). 
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Wildlife

Rohr et al. (2008) exposed green frog (Rana clamitans) tadpoles to environmentally relevant 

concentrations of one of four pesticides (atrazine, glyphosate, carbaryl, or malathion) for 

one week, then half were exposed to a trematode parasite, (Echinostoma trivolvis), for 24 

hours, and then all were returned to the same pesticide exposure for an additional week. The 

authors reported that exposure to each of the pesticides increased the percentage of encysted 

cercariae compared with tadpoles exposed to a solvent control. They stated that the increased 

susceptibility may be a result of immunosuppression. Additional studies have shown sublethal 

and other effects on receptor organisms exposed to low levels of pesticides. For example, Hayes 

et al. (2003) showed that atrazine can induce hermaphrodism in leopard frogs (Rana pipiens), and 

Blakley et al. (1999) showed immunotoxicological effects of pesticides to a variety of organisms. 

Organochlorine Pesticides

Toxicological effects of organochlorine pesticides and metabolites upon Chesapeake waterbirds 

occurred principally through impairment of reproduction rather than through altered survival 

of adults (Heinz and Wiemeyer 1991; Blus 1996). As described in various reviews (Ohlendorf and 

Fleming 1988; Heinz and Wiemeyer 1991; Rattner and McGowan 2007), concentrations of the DDT 

metabolite p,p’-DDE in osprey (Pandion haliaetus) and bald eagle (Haliaeetus leucocephalus) eggs 

in the Chesapeake were among the highest in the nation and appear to have been the principal 

cause of the population declines that were observed in these species from the 1950’s through 

1970’s in this region (Wiemeyer et al. 1984, 1988, 1993). Many other organochlorine pesticides 

and metabolites were also detected in eggs and young during this era (e.g., dieldrin, endrin, 

heptachlor epoxide, chlordane, oxychlordane, mirex, toxaphene). Since the phase out of some 

organochlorine pesticides in the early 1970’s, Chesapeake Bay osprey and bald eagle populations 

gradually began to rebound and currently populations are the greatest since they began to be 

systematically monitored (Watts et al. 2004, 2007).

Since 2000, data are available for only a few addled eggs of bald eagles from the Chesapeake. 

A single addled egg collected at Aberdeen Proving Ground in Harford County Maryland in 2008 

contained p,p'-DDE at a concentration of 9.46 ppm wet weight, and chlordane constituents and 

metabolites (e.g., oxychlordane, cis- and trans-nonachlor, cis- and trans-chlordane) totaling 3.34 

ppm (Mojica and Watts 2008). Notably, p,p'-DDE exceeded 5.5 ppm, and threshold concentrations 

associated with embryo lethality (Elliott and Harris 2002; Henny and Elliott 2007). Two addled 

eggs collected from the Indian Head Naval Support Facility in Charles County Maryland in 2008 
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and 2009 contained p,p'-DDE at concentrations of 3.77 and 5.61 ppm, chlordane constituents and 

metabolites at concentrations of 2.12 and 4.34 ppm (Mojica and Watts 2011). Organochlorine 

pesticides and metabolites in blood of eagle nestlings were also quantified in 58 blood samples 

of nestling eagles from these same sites (Mojica and Watts 2008, 2011). The sum of DDT and 

metabolites (< 0.1 ppm wet weight) and the sum of chlordane constituents ( < 0.02 ppm) were 

below the thresholds for toxicity (Elliott and Harris 2002; Henny and Elliott 2007).

In 2000 and 2001, a large-scale study was undertaken in which osprey eggs were collected from 

nests in these regions of concern (Baltimore Harbor, Anacostia River, Elizabeth River) and nearby 

tributaries, and from a reference area (South, West and Rhode Rivers near Annapolis) (Rattner et 

al. 2004). Of 27 organochlorine pesticides and metabolites quantified, 17 or more were detected 

in half of the sample eggs. Concentrations of p,p’-DDE ranged from 0.26-1.92 ppm wet weight in 

the regions of concerns compared to 0.24-0.98 ppm in the reference area. Over half of the eggs 

sampled from the Anacostia and middle Potomac Rivers had p,p’-DDE concentrations within the 

95% confidence interval (1.2 to 3.0 ppm) associated with 10% percent eggshell thinning in osprey 

eggs, but no samples exceeded the value (>2.6 ppm) associated with population instability 

(production of 1 fledgling per active nest) (Wiemeyer et al. 1988). The concentration of p,p’-DDD 

ranged up to 0.24 ppm, while other compounds (e.g., dieldrin, endrin, heptachlor epoxide, αand γ-chlordane, oxychlordane, cis- and trans-nonachlor, mirex, HCB) were well-below 0.1 ppm, 

and seemingly below reproductive toxic thresholds associated with in bird eggs. Chlordecone 

(also known as Kepone®) an organochlorine insecticide had been found at concentrations of up 

to 5.0 ppm wet weight in osprey eggs collected from Chesapeake tributaries in Virginia in 1977 

(Wiemeyer et al. 1988), but was not detected in a subset of osprey eggs from the Elizabeth River 

in 2001 (Rattner et al. 2004).

Between 1993 and 1999, post-term and addled peregrine falcon (Falco peregrinus) eggs were 

collected from the Chesapeake region (e.g., Chincoteague and Martin National Wildlife Refuges, 

Clay, Cobb, Fisherman, Metompkin, South Marsh, and Wallops Islands, Baltimore, Wachapreague, 

and Fort Eustis) as part of a mid-Atlantic states monitoring effort (Clark et al. 2009). 

Concentrations of p,p’-DDE ranged from 0.94 to 5.05 ppm wet weight. Other organochlorine 

pesticides in these egg samples were generally low (e.g., β-hexachlorocyclohexane <0.01 ppm, 

oxychlordane <0.5 ppm, cis-nonachlor <0.25 ppm, trans-nonachlor <0.3 ppm, HCB <0.02 ppm, 

hepatochlor epoxide <0.3 ppm, and o,p’-DDD, p,p’-DDD, o,p’-DDT, and p,p’-DDT <0.02 ppm), 
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with the possible exception of dieldrin (range 0.05 to 1.31 ppm). Despite the presence of these 

organochlorine pesticides, 11 of 14 clutches that were followed in detail were successful in 

producing young. In another study, 22 peregrine eggs that failed to hatch in 2001 and 2002 

from the Chesapeake region had p,p’-DDE concentrations ranging from 1.38 to 12.4 ppm wet 

weight (Potter et al. 2009). Notably, concentrations in two of the eggs exceeded 10 ppm, and thus 

approached the threshold for 20% eggshell thinning and impaired reproduction. 

A study assessing reproduction and contaminant exposure of the Baltimore Harbor blackcrowned night heron (Nycticorax nictocorax) colony was undertaken in 1998 (Rattner et al. 2001). 

Based on analyses of a large number of sample eggs, concentrations of p,p’-DDE (0.023-1.292 

ppm wet weight) and dieldrin (not detected to 0.262 ppm) were clearly lower than observed in 

1991, and levels of other organochlorine pesticides were also quite modest. Pesticides in nightheron eggs concurrently collected from Holland Island, a remote reference colony in the southern 

Chesapeake, were but a fraction of those observed in Baltimore Harbor. There was neither 

eggshell thinning nor impaired reproduction in night–herons that could be associated with DDT 

exposure in this study.

Two studies examined organochlorine pesticides in common tern (Sterna hirundo) eggs. In 

1997, 10 three-egg clutches were collected from Bodkin Island, and organochlorine pesticide 

concentrations (e.g., DDT and metabolites, chlordane and metabolites, dieldrin, endrin, heptachlor 

epoxide and mirex) were below the detection limit of 0.061 ppm dry weight (French et al. 2001). In 

a 2010 collection on Poplar Island, eggs of 6 terns contained low levels (<0.08 ppm wet weight) of 

p,p’-DDE, cis- and trans-nonachlor, mirex and hexachlorobenzene (Rattner et al. 2010).

Organophosphorus and Carbamate Pesticides

Use of organophosphorus and carbamate pesticide in the United States expanded greatly 

following the ban of many organochlorine pesticides in the 1970’s. These compounds or 

their activated metabolites inhibit the enzyme acetylcholinesterase, thereby disrupting 

neurotransmitter processes in the central nervous system and normal neural functioning of the 

sensory, integrative, and neuromuscular systems (Ballantyne and Marrs 1992). Definitive evidence 

of poisoning in wildlife usually entails detection of the parent compound in the digestive tract 

and substantial inhibition of brain acetylcholinesterase activity (Hill 2003).
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For the period of 1988 through 2011, the Contaminant Exposure and Effects-Terrestrial 

Vertebrates database (Rattner et al. 2005; Rattner and McGowan 2007) contained over 130 records 

documenting at least 50 incidents of unintentional bird poisonings in the Chesapeake Bay region 

that involved anticholinesterase pesticides, and most of these incidents occurred in agricultural 

settings. Diagnoses are derived from inhibition of brain cholinesterase activity, circumstantial 

evidence of use in nearby fields, and in a few instances, detection of parent compound in the 

gastrointestinal tract. These compounds do not accumulate appreciably in tissue of higher 

vertebrates, so tissue residue data are not available. These die-offs involved six orders of birds, 

most frequently passerines, birds of prey, and waterfowl. These incidents were principally caused 

by carbofuran exposure, and affected many different species of birds, including at least one bald 

eagle (Stinson et al. 1994). Use of granular formulations of this carbamate was restricted by the 

USEPA in 1994 due to bird die-offs. The remaining incidents involved chlorpyrifos, diazinon, 

fenthion, phorate, parathion and the carbamate oxamyl. 

With the exception of aforementioned pesticides, there appear to be no residue data for the 

vast majority of some other compounds (e.g. fungicides, some insecticides, rodenticides, etc.) in 

free-ranging Chesapeake Bay wildlife (amphibians, reptiles, birds and mammals). This apparent 

data gap reflects the absence of routine monitoring for exposure to newer chemical groups 

of compounds, their shorter environmental half-lives, rapid metabolism, lower potential to 

bioaccumulate, paucity of analytical laboratories that routinely quantify newer compounds, and 

the implementation metrics that associate both exposure and declines in health and condition 

of wildlife.

It would appear that the frequency of outright mortality of wildlife from pesticide exposure in 

the Chesapeake region has declined considerably since the organochlorine pesticide era. There 

are probably still anticholinesterase poisoning incidents, but their frequency and magnitude 

do not seem to be exerting population level effects. Of the newer groups of pesticides, the 

ecological consequences of chronic low level exposure and potential adverse effects are 

inadequately known.

Conclusions

Because of the diverse uses, modes of action, and chemical characteristics of pesticides, the 

available data on environmental occurrence and related toxicology range from minimal to 

extensive. In addition, some pesticides are included in current monitoring programs, whereas 
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many others are not. On the basis of data collected within the Chesapeake Bay watershed, 

herbicides as a group are known or are suspected to occur throughout the watershed. The 

likelihood that current and future land uses will include regular and widespread applications 

of herbicides in agricultural, as well as other areas where lawns and other plantings are 

maintained increases their potential to be widely distributed in the environment. Specifically, 

atrazine and its degradates are widely distributed throughout the agricultural parts of the 

watershed, and recent studies have shown the potential for atrazine to cause adverse effects 

at environmentally relevant concentrations. Insecticides and fungicides, though potentially 

important environmental contaminants as a result of their potential effects on non-targeted 

insects, fungi, and other organisms (e.g., broad-spectrum insecticides, systemic fungicides), 

have not been monitored as broadly or as consistently as herbicides; however, some of the more 

recent studies indicate the potential for a range of effects related to insecticides and fungicides at 

environmentally relevant concentrations. Therefore the occurrence of herbicides, and specifically 

atrazine and its degradates, is identified as widespread. Insecticides such as aldrin, chlordane, 

dieldrin, DDE/DDT, heptachlor epoxide, mirex, and their degradates occur in localized areas at 

levels concentrations high enough to indicate that adverse effects are possible. Some legacy 

pesticides persist for long periods of time locally as well as regionally in areas of current and past 

usage. Data and research gaps exist for many pesticides including some current-use and some 

legacy pesticides, and consequently the extent and severity remains uncertain and cannot be 

evaluated at this time. For example, the potential sublethal effects of low concentrations of many 

pesticides (and degradates) and their mixtures (including adjuvants, etc.) in the environment is 

poorly understood.

2.6 Pharmaceuticals

Abstract

Pharmaceuticals are chemicals used in the diagnosis, mitigation, treatment, cure, and prevention 

of disease in humans and other animals. There is potential widespread extent of occurrence 

of some pharmaceutical compounds in the environment due to the abundant human (e.g., 

municipal wastewater-treatment plant discharges) and animal waste (e.g., livestock feeding and 

related activities) sources in the watershed and their pathways to the environment. The few 

occurrence studies in the watershed as well as many studies elsewhere in the United States and 

across the globe have documented pharmaceuticals as environmental contaminants. In addition, 

although effects on receptor organisms are still poorly understood, several studies have shown 

biological uptake in a wide variety of exposed aquatic and other species, as well as a range of 
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sublethal effects of some compounds at environmentally relevant concentrations (e.g. synthetic 

hormones, some antidepressants, diclofenac, ciprofloxacin and sulfamethoxazole). 

Although there is evidence these compounds are of potential concern as toxic contaminants in 

the Bay watershed, major data gaps and lack of understanding of effects limit assessment. For 

these reasons, extent and severity of pharmaceuticals are identified as uncertain.

Background

Pharmaceutical compounds have been considered potential environmental contaminants 

since the 1960’s (Richardson and Bowron, 1985) and their occurrence in rivers, groundwater, 

and untreated drinking water sources of the United States has been documented for many 

years (Kolpin et al. 2002; Focazio et al. 2008). Although broader definitions of pharmaceuticals 

include herbal preparations, nutraceuticals, cosmeceuticals, and other remedies, this section 

focuses on the conventional prescription and nonprescription pharmaceuticals as noted in 

the Food, Drug and Cosmetic Act (FD&C Act, sec. 201(g)(1)). Excipients, which are additives in 

pharmaceuticals that give a drug its form, control its release into the bloodstream, facilitate its 

transport to targeted sites in the body, preserve its shelf life, and enhance its taste, are beyond 

the scope of this report but could be important compounds for future research on environmental 

contaminants generally. This report focuses only on the active pharmaceutical ingredients. In 

addition, biological drugs or so-called “biopharmaceuticals” such as vaccines that are typically 

administered by injection in the doctor’s office are not considered here. The chemically derived 

prescription and nonprescription drugs covered in this report range from popular over-thecounter analgesics to rigorously controlled chemotherapy or other drugs. 

Pharmaceuticals are also commonly used to treat domestic pets as well as livestock. Livestock 

usages include growth promotion, estrus modulation, and various purposes that are not common 

goals of human pharmaceutical usage. Many human-use pharmaceuticals are also used as 

veterinary pharmaceuticals, though many veterinary pharmaceuticals are restricted to animal 

use only. In addition, many over-the-counter pharmaceuticals are used for domestic pets and 

livestock. Veterinary drugs are administered for a variety of reasons, including nervous system 

(anesthetic), digestive tract and metabolism (antacid), antiparasitic (anthelmintic, antibacterial, 

antimicrobial, antiseptic, astringent, anti-infective), respiratory (antihistamine), steroidal and 

nonsteroidal anti-inflammatory, respiratory (bronchodilator), genitourinary (diuretic, emetic), 

estrus synchronization (sex hormones), nervous system (sedative, tranquilizer), and growth 

promotant purposes.
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According to 1995 Toxics Release Inventory (TRI) data, facilities released (discharged to the 

air, water, or land without treatment) and transferred (shipped off-site) a total of 177 million 

pounds of contaminants, made up of 104 different chemicals (USEPA 1997). This amount 

represents about 3 percent of the 5.7 billion pounds of TRI chemicals released and transferred 

by all manufacturers in that year. In comparison, the chemical industry as a whole produced 

1.7 billion pounds that year, accounting for about 30 percent of all releases and transfers. Of 

the pharmaceutical industry’s TRI releases, 57% go to the atmosphere, 25% go to underground 

injection, 17% go to surface waters, and 1% go to the land. The reported chemical releases from 

pharmaceutical manufacturing facilities are not commonly active pharmaceutical ingredients 

(APIs), but rather volatile organic compounds (VOCs) (e.g., solvents) and other chemicals 

used in the manufacturing process itself. It is not possible at this time to summarize the API 

releases. USEPA defines high-production-volume chemicals as those produced or imported at 

quantities exceeding 1 million pounds per year; therefore, some pharmaceuticals and classes of 

pharmaceuticals are considered high-production-volume chemicals whereas many are not. In 

addition effluents associated with pharmaceutical manufacturing plants have been shown to 

be important localized sources of active ingredients in receiving streams (Phillips et al. 2010). 

Consequently national estimates of production volume may not be an adequate indicator of the 

potential for a pharmaceutical to affect ecological receptors. 

Approximately half of the U.S. population takes at least one prescription drug during any given 

month, with concurrent use of several drugs being common in the older populations (Gu et al. 

2010). Therefore, as more than 17 million people live in the Chesapeake Bay watershed, about 

8.5 million prescription drugs are consumed in any given month, or about 102 million per year. 

Nonprescription drug use is more difficult to track for many types of drugs (e.g., analgesics, 

antihistamines, etc.). It is beyond the scope of this report to assess and summarize the mass of 

pharmaceuticals manufactured and released in the Chesapeake Bay watershed; however, it is 

clear that they are manufactured in large quantities throughout the United States, are consumed 

within the Bay watershed by the majority of the population and, therefore, are being released 

to the environment in human and animal wastes and waste-management activities in many 

locations throughout the watershed. 

As a group, pharmaceuticals are complex molecules with various acidic and basic functionalities; 

they exhibit a range of neutral, cationic, anionic, or zwitterionic behaviors in the environment, are 

of high molecular weight (commonly ranging from 300 to 1,000 daltons), and are biotransformed 
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in the body, commonly resulting in changes in their physical and chemical properties upon 

excretion (Kummerer 2004). For example, Sarmah et al. (2006) listed selected antibiotics 

and showed the range of physical and chemical properties that exists for this one group of 

pharmaceuticals. Once in the environment, APIs can further undergo a range of potential 

transformations including photolysis, bacterially remediated degradation, hydrolysis, and dilution. 

Generally, APIs are biologically active and are mobile because they have high water solubilities in 

relation to molecular weight (Kummerer 2004). 

Pharmaceuticals are associated with point as well as nonpoint sources and pathways to the 

environment. In sewered areas, pharmaceuticals are transported from human excretion or 

disposal in residential settings to municipal wastewater-treatment plants (WWTPs). In this way, 

WWTPs represent centralized locations where various residential, hospital, manufacturing, and 

other potential pharmaceutical wastes are accumulated before being discharged as a point 

source of mixtures of APIs and their metabolites to receiving water bodies. Most households in 

urban areas of the U.S. are connected to municipal sewers, which often include hospitals and can 

include pharmaceutical manufacturing facilities within the same sewersheds as well. Although 

WWTPs are not designed to remove APIs or their metabolites, studies have shown that existing 

treatment processes can decrease the concentrations of some parent compounds, in some cases 

to below detection (Phillips et al. 2012; Drewes 2007) before being discharged. Barber et al. (2012) 

documented a decrease in estrogenicity of wastewater effluents and an associated decrease in 

intersex conditions of exposed fish due to an upgrade in treatment. Other APIs, however, are 

discharged relatively unchanged because of their ability to resist the type of treatment typically 

found in municipal WWTPs. Other studies (Benotti and Brownawell 2007, Stackelberg et al. 2007) 

have shown that drinking-water treatment is effective in decreasing concentrations of parent 

pharmaceutical compounds in many cases; in some limited cases, however, the pharmaceutical 

compounds are recalcitrant. Considering the amount of pharmaceuticals manufactured and 

prescribed as well as factors controlling the removal or dilution of pharmaceuticals as they 

are transported from wastewater treatment plants to receiving streams, Anderson et al. (2004) 

developed an empirical model to determine the potential environmental concentrations of 

targeted pharmaceuticals released to aquatic environments via wastewater treatment plants 

throughout the United States. Schwab et al. (2005) used the model to predict concentrations of 

26 human-use pharmaceuticals in U.S. surface waters and concluded that “no appreciable human 

health risk exists.” Recently the model was revised to include predictions of pharmaceutical 

concentrations in biosolids generated in wastewater treatment plants (Cunningham et al. 2012). 
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Other sources of human-waste-related APIs include point-source leachate from onsite sewage 

disposal such as septic tanks (Conn et al. 2006). Phillips et al. (2010), in an analysis of effluent 

from WWTPs in New York State (outside the Chesapeake Bay watershed) with a focus on 

sewersheds that contain pharmaceutical manufacturing facilities, conclude that manufacturing 

practices can result in pharmaceuticals concentrations from 10 to 1,000 times higher than 

those typically found in WWTP effluents. Point sources of wastewater effluent are the primary 

source pathways of human-use pharmaceuticals to the environment in the urbanized areas of 

Chesapeake Bay watershed, as they are elsewhere in the United States. Therefore, although some 

generalizations can be made regarding the expected environmental occurrence and persistence 

of some individual or classes of pharmaceuticals, accurate understanding of the ultimate fate of 

these compounds and their severity in the environment requires site- and compound-specific 

assessments. 

Pharmaceuticals are also associated with nonpoint sources and pathways to the environment 

where agricultural land uses, particularly animal feeding operations, are present (Sarmah et 

al. 2006). Sources of APIs include feed or drinking-water additives, direct injections, implants, 

drenches, and pastes. The use and length of treatment and whether the drug is delivered to 

an individual animal or to a herd or flock determine, in part, how a specific drug is ultimately 

reaches the environment (Sarmah et al. 2006). Potential pathways to the environment reflect 

overall manure management practices and include composting and/or application of biosolids/

manures on crop or other lands, deposition by animals in liquid or solid wastes excreted directly 

into surface-water bodies, and releases of liquid animal wastes from lagoons or other holding 

tanks, as well as the local hydrologic condition. Therefore, the potential agricultural sources and 

the pathways depend, in large part, on how animals and their wastes are managed. In a recent 

summary of the literature on animal manures, the USEPA (2000c) showed that the amount of 

manure per acre of farmland in states within the Chesapeake Bay watershed is among the highest 

in the Nation. 

As a group, pharmaceuticals represent a wide array of potentially biologically active compounds 

with a variety of potential mechanisms of action. Additionally, although these compounds are 

designed for specific pharmacological purposes, they also can exhibit a range of unintended 

effects on target as well as nontarget species at therapeutic doses. The toxicity of therapeutic 

doses of pharmaceuticals is well understood because these chemicals are intentionally 

ingested and because the FDA registration process requires it (U.S. Department of Health and 
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Human Services 2007). The effects of pharmaceuticals on aquatic or terrestrial organisms at 

environmentally relevant concentrations are less well understood. Recent research is focused on 

the potential effects on organisms that are exposed to low concentrations of pharmaceuticals 

in the environment. Currently, no aquatic-life or related water-quality or sediment-related 

benchmarks exist for pharmaceuticals; however, several studies have documented a range of 

lethal and sublethal effects on receptor organisms in laboratory, mesocosm, and field settings. 

Swan et al. (2006) was one of the first studies to link environmental exposures of a pharmaceutical 

to an adverse ecological effect. They found that vultures (Gyps spp) in South Asia feeding on 

cattle carcasses with high levels of diclofenac (a veterinary pharmaceutical) died of renal failure 

in numbers that threatened their extinction. Since then an increasing number of studies, 

many focused on aquatic organisms, have been undertaken. In addition, there continue to be 

more efforts to demonstrate and quantify biological uptake of environmental pharmaceuticals 

in various species and environmental source pathways. For example, Kinney et al. (2012) 

documented bioaccumulation of wastewater indicator compounds (commonly indicative 

of sources of environmental pharmaceuticals) in earthworms exposed to biosolids. Other 

studies, such as Ramirez et al. (2009), Bringolf et al. (2010), and Schultz et al. (2011), are providing 

baseline data and information on biological uptake and effects of targeted pharmaceuticals 

by various species living in municipal wastewater-affected streams. In related work, Li et al. 

(2012) documented a range of antibiotics including quinolones, macrolides, and sulfonamides 

in mollusks collected from the Bohai Sea of China. Kidd et al. (2007) documented populationlevel declines in native fathead minnows in a Canadian lake exposed to environmentally 

relevant concentrations of ethinyl estradiol (synthetic estrogen). Other pharmaceuticals such as 

antidepressants have been shown to affect the behavior of fish living downstream of wastewater 

treatment plants (Schultz et al. 2011). Thus, an increasing number of studies are showing 

effects of pharmaceuticals on various organisms and ecosystems at environmentally relevant 

concentrations.

In this report, “pharmaceutical” (or synthetic) hormones are considered separately from naturally 

occurring (or biogenic) hormones (see the “Biogenic Hormones” section of this report). Although 

their modes of action are similar, their potencies, sources, and other factors are different. For 

simplicity, therefore, these two major groups of hormones have been separated. A third group 

with similar hormone-related modes of action includes other chemicals that have been shown 

to interfere with normal endocrine functioning through antagonistic or synergistic effects on 

hormone receptors. These other known, or suspected, “endocrine-disrupting” chemicals, which 
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include some pesticides, surfactants (e.g., alkylphenol ethoxylates), ingredients in plastics (e.g., 

bisphenol A), and flame-retardant mixtures (e.g., polybrominated diphenyl ethers), are discussed 

in other sections of this report. 

Although representative studies conducted in the Chesapeake Bay watershed are highlighted 

in this section, several studies conducted outside the watershed are included for illustrative 

purposes. A short list of representative laboratory, mesocosm, and field studies that have 

linked various categories of environmental exposures to three major classes of pharmaceuticals 

(synthetic hormones, antidepressants, and antibiotics) to a range of adverse effects on biological 

receptors is included. This list is meant to illustrate important effects of these pharmaceuticals on 

fish and wildlife health only and does not represent a complete literature review.

Water

Human-health or ecologically based water-quality standards are not common for pharmaceuticals 

and they are not regulated water-quality constituents, consequently, they are not commonly 

monitored in the environment. However, a few research-based datasets for the Chesapeake Bay 

watershed and a few for the Nation as a whole are available. State integrated assessments did 

not include pharmaceuticals, but the few studies that assessed pharmaceuticals as environmental 

contaminants in the Bay watershed are summarized below. Several studies that provide a 

national perspective are summarized as well.

Glassmeyer et al. (2008) reviewed and summarized the available literature on the occurrence of 

pharmaceuticals and personal-care products in the environment worldwide and showed that 

most of the current papers originated in the United States. Although the individual studies 

summarized in the report included a variety of compounds with different analytical methods, 

detection limits, and research objectives, it is clear that some groups of pharmaceuticals tend 

to be detected more frequently than others in the aquatic environments sampled (including 

groundwater, surface water, WWTP influents, and WWTP effluents, as well as biosolids and 

streambed sediments). Sixty-seven compounds were found in wastewater influents, 105 were 

detected in wastewater effluents, and 22 were present in biosolids. One-hundred twenty-four 

compounds were detected in surface waters, 35 were detected in groundwater, and 4 were 

found in bed sediments. Of the 126 compounds measured in at least one matrix, 18, including 

analgesics, antibiotics, antihyperlipidemics, an antiepileptic, and other compounds such as 

galaxolide and tonalide, which are used as fragrances in personal-care products, were detected 
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in more than 10 different studies. In another national study by the USGS (Focazio et al. 2008), 25 

groundwater and 49 surface-water sites that are sources of drinking water were sampled for the 

occurrence of 100 targeted chemicals including pharmaceuticals, personal-care products, and 

other organic wastewater compounds. Forty percent of the 36 pharmaceuticals analyzed were 

detected at least once, and carbamazepine was the most commonly detected pharmaceutical in 

both surface water and groundwater. 

Phelan and Miller (2010) collected samples from 23 stream sites in Rock Creek National Park 

(Washington, DC) selected on the basis of warm water thermal infra-red images. The samples 

were analyzed for a suite of nine human-use pharmaceuticals. Butalbital and oxycodone were 

detected but concentrations could not be quantified as a result of analytical uncertainties. 

Although the sources of the warm water were not specifically linked to pharmaceutical 

detections, they nevertheless indicate the presence of leaking municipal wastewater pipes or 

other anthropogenic influences on the water. 

Loper et al. (2007) present environmental and quality-control data from analyses of water samples 

from 11 streams and 6 wells in south-central Pennsylvania for 46 pharmaceuticals (31 of which 

were human and veterinary antibiotics). Five of the streams received municipal wastewater 

and six received runoff from agricultural areas dominated by animal-feeding operations. For all 

11 streams, samples were collected at locations upstream and downstream from the municipal 

effluents or animal-feeding operations. All six wells were in agricultural settings. For the stream 

samples, 24 pharmaceuticals (the authors of this report included caffeine and its metabolite as 

pharmaceutical compounds) were detected at least once. Of the pharmaceuticals detected, 11 

were antibiotics. Seventy-eight percent of all detections were in samples collected downstream 

from municipal-wastewater effluents. The maximum concentrations of compounds other than 

caffeine in the wastewater effluents include azithromycin (1.65 μg/L; antibiotic), sulfamethoxazole 

(1.34 μg/L; antibiotic), carbamazepine (0.516 μg/L; antiepileptic), ofloxacin (0.329 μg/L; antibiotic), 

ibuprofen (0.277 μg/L; nonsteroidal anti-inflammatory), and trimethoprim (0.256 μg/L; antibiotic). 

Concentrations and detections were much lower in streams receiving runoff from animalfeeding operations than in those receiving municipal-wastewater effluents during baseflow. 

The pharmaceuticals detected were acetaminophen, carbamazepine, diphenhydramine, 

oxytetracycline, sulfadimethoxine, sulfamethoxazole, and tylosin. The maximum concentration 

for all pharmaceuticals was 0.157 μg/L. In samples from wells used to supply livestock, 3 of the 24 

pharmaceutical compounds were detected—diphenhydramine, tylosin, and sulfamethoxazole. 
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There were five detections in all the well samples. The maximum concentration for cotinine 

detected in well water was 0.024 μg/L (estimated).

Alvarez et al. (2008) used passive samplers to assess the occurrence of a limited number of 

pharmaceuticals and other wastewater indicator compounds at nine locations in the Shenandoah 

River Basin and two in the James River Basin, Virginia. Detections included codeine (a narcotic 

analgesic) and carbamazepine (an anticonvulsant drug), each of which was detected at several 

sites. Trimethoprim, an antibi¬otic commonly prescribed in tandem with sulfamethoxazole, was 

detected at 8 of 10 sites; however, sulfamethoxazole was detected only at 1 of these sites. The 

antidepressant venlafaxine, currently the 13th most prescribed drug in the United States was 

detected at several sites.

Arikan et al. (2008) collected water samples from 22 stream sites in the Choptank River, Maryland, 

including 7 sites on the mainstem and 15 on tributaries during baseflow. Samples were 

collected in all four seasons (April, June, September, and December) in a reconnaissance effort 

to assess the occurrence of key tetracycline and sulfur classes of antibiotic compounds. Of the 

sulfonamide class of antibiotics, sulfamethoxazole (19% detection) and sulfadimethoxine (12% 

detection) were detected at the mainstem stations at maximum concentrations of 0.002 and 

0.003 μg/L, respectively. The most frequently detected compounds at the tributary stations 

were sulfamethoxazole (5% detection) and sulfadimethoxine (14% detection), with maximum 

concentrations of 0.007 and 0.009 μg/L, respectively. Of the tetracycline class of antibiotics, 

chlortetracycline (19% detection) and oxytetracycline (15% detection) were the most frequently 

detected of the tetracycline group of antibiotics at the mainstem stations, with maximum 

concentrations of 0.034 and 0.047 μg/L, respectively. Tetracycline and doxycycline were detected 

only once (4% detection) at the mainstem stations, at concentrations of 0.005 and 0.020 μg/L, 

respectively. Chlortetracycline (21% detection) and oxytetracycline (18% detection) were 

the most frequently detected tetracycline antibiotics at the tributary stations, with maximum 

concentrations of 0.180 and 0.084 μg/L, respectively. Tetracycline and doxycycline were 

detected at 5% of the tributary stations, with maximum concentrations of 0.003 and 0.146 μg/L, 

respectively. In this limited dataset, seasonal trends were not clear; however, the authors report 

more samples with detections for antibiotics in December (chlortetracycline was detected in 14 

of 19 samples collected in December) than in the other seasons. The highest concentration for all 

sulfonamide antibiotics was 0.694 μg/L for sulfamerazine in a sample collected in April.
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Barnes et al. (2008) reported datasets used in a national reconnaissance of pharmaceuticals 

and other organic wastewater compounds in untreated sources of drinking water in 25 states 

throughout the United States, including three states in the Chesapeake Bay watershed (six wells 

in Pennsylvania, three wells in West Virginia, and four surface-water intakes in Virginia). Results 

indicated that sulfamethoxazole (antibiotic) was detected in a Pennsylvania well and none of 

the compounds analyzed for were detected in the West Virginia wells. Sarafloxacin (antibiotic) 

and carbamazepine (antiepileptic) were each detected once in samples from Virginia surfacewater sites. 

The USGS is conducting ongoing research throughout the Chesapeake Bay watershed to 

determine whether fish health is related to chemical exposure. Initial research included 

the sampling of water and bed sediment at seven active smallmouth bass nesting sites 

during spawning (Kolpin et al. 2013). During this study, 14 prescription and nonprescription 

pharmaceuticals (including caffeine and its metabolite as well as cotinine, a metabolite of 

nicotine) were detected in the water samples collected. The two most frequently detected 

compounds were caffeine (86%, stimulant) and iso-chlorotetracycline (71%, antibiotic degradate). 

Other pharmaceuticals detected include acetaminophen (29%, analgesic), azithromycin (14%, 

antibiotic), carbamazepine (29%, antiepileptic), diphenhydramine (14%, antihistamine), epiiso-chlorotetracycline (15%, antibiotic degradate), epi-tetracycline (29%, antibiotic degradate), 

oxytetracycline (43%, antibiotic degradate), sulfamethazine (29%, antibiotic), sulfamethoxazole 

(29%, antibiotic), sulfathiazole (14%, antibiotic), tetracycline (57%, antibiotic), and tylosin (14%, 

antibiotic). 

Sediment

Very few sediment analyses have been published for samples collected in the Chesapeake Bay 

watershed; however, Glassmeyer et al. (2008), in a national summary of pharmaceutical and 

other organic wastewater data, note that whereas the concentrations of pharmaceuticals in 

liquid effluents typically are in the range of low parts per billion to high parts per trillion (low 

micrograms per liter (μg/L) to high nanograms per liter (ng/L)), they are found in biosolids at parts 

per million (milligrams per kilogram (mg/kg)) concentrations. These values reflect, in part, the 

affinity of many of these compounds to attach to solid particles and the manner in which solid 

wastes from municipal wastewater-treatment plants and other sources are managed. 
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Arikan et al. (2008) analyzed bed sediment from four sites on the Choptank River. 

Chlortetracycline (100% detection) and sulfamethoxazole (75% detection) were the most 

frequently detected of the antibiotics. The maximum concentrations were 10.0 and 0.15 ppb 

dry weight, respectively. Sulfamethazine was detected at one site at a concentration of 0.82 ppb 

dry weight. Chlortetracycline and sulfamethoxazole were also the most frequently detected 

antibiotics in water samples at these sites. Oxytetracycline, the second most frequently detected 

compound in water samples from these sites, was not detected in bed-sediment samples. Kolpin 

et al. (2013) also analyzed bed sediment for a limited number of compounds. 

Fish and Wildlife

No fish tissue analyses for pharmaceuticals from the Bay watershed were identified. However, 

several studies conducted outside the watershed have shown that pharmaceuticals such as 

antidepressants can be detected in the brain tissue of fish (Schultz et al. 2011) where the fish have 

been exposed to wastewater effluents in aquatic environments.

Pharmaceutical (Synthetic) Hormones

Much of the environmental research regarding pharmaceutical hormones has focused on the 

synthetic sex hormones such as ethinyl estradiol, the active ingredient in human birth control 

pills. Leet et al. (2011) summarized the state of the science on the role of exposure to estrogens 

and androgens in sexual differentiation during early life stages of fish. Other hormonally active 

pharmaceuticals such as those used for thyroid-related therapies, corticoid-related therapies, 

and others are beginning to receive more attention as potential environmental contaminants; 

however environmentally relevant studies of these other pharmaceutical hormone groups are not 

common (Kugathas and Sumpter 2011). 

Some of the earliest work linking reproductive endocrine disruption effects and potential 

environmental exposures to pharmaceuticals (ethinyl estradiol and diethylstilbestrol) was 

demonstrated in Japanese quail (Coturnix japonica) by Halldin et al. (1999). The study showed 

embryonic exposure to these synthetic estrogens can affect sexual differentiation and cause 

reproductive impairment. Research on synthetic and other exogenous hormones has improved 

the understanding of the effects of hormones in the environment regardless of whether the 

hormones are derived from natural sources or are synthetic. Additional research has been 

performed on other chemicals that are known to interfere with normal hormone-receptor 

functioning in vertebrates. For example, the effect of embryonic diethylstilbesterol (DES, a 
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synthetic estrogen) exposure demonstrated that estrogenic chemicals affect the differentiation of 

estrogen target organs (Colborn et al. 1993). Data on androgen- and thyroid-active compounds 

reveal similar deleterious potential for long-term effects from these compounds through different 

mechanisms of action (Custer et al. 1999; Chen et al. 2008; Zoeller 2008).

 

Other research has shown that in some cases exposure to estrogens at environmentally relevant 

concentrations can result in immunosuppressive effects (Milla et al. 2011). For example, 

Roberston et al. (2010) have shown that expression of a protein critical to immune suppression 

(hepcidin) can be disrupted in fish exposed to exogenous estrogens at environmentally relevant 

concentrations (also see the “Biogenic Hormones” section of this report). Cubero-Leon et al. 

(2010) have documented that natural (estradiol) as well as synthetic (ethinyl estradiol) estrogen 

can disrupt serotonin receptor functioning as well as mRNA expression levels of cyclooxygenase 

(an enzyme critical to production of prostaglandins) in the marine bivalve Mytilus edulus. More 

recently, Kidd et al. (2007) dosed a Canadian lake with environmentally relevant concentrations 

(i.e., concentrations detected by other studies in environmental waters) of synthetic estrogen 

(17α-ethinyl estradiol) over a 3-year period. Population-level effects including intersex and severe 

reproductive failure were observed in fathead minnows (Pimephales promelas) from the dosed lake 

but not in those from two control lakes. Populations returned to normal when the contaminant 

dosing was discontinued. Other studies have shown that total estrogenicity of common mixtures 

of contaminants (composed of natural, pharmaceutical, and other hormonally active compounds) 

such as those commonly found in municipal wastewater effluents can cause intersex in exposed 

fish (Vajda et al. 2011). Critically, once the wastewater-treatment plant was updated with 

treatment technologies capable of removing these estrogenic compounds, the intersex conditions 

in fish were no longer evident (Barber et al. 2012). 

Pharmaceutical Antidepressants

Schultz et al. (2011) exposed fathead minnows to environmentally relevant concentrations of 

antidepressants (bupropion, fluoxetine, sertraline, and venlafaxine) and observed anatomical 

and physiological effects, including the presence of intersex biomarkers such as elevated plasma 

vitellogenin in male fish as well as severe effects on male secondary sexual characteristics. Some 

exposures to venlafaxine (305 ng/L) and sertraline (5.2 ng/L) resulted in mortality. Painter et 

al. (2009) showed that predator avoidance behaviors were adversely affected in larval fathead 

minnows exposed to these antidepressants at environmentally relevant concentrations. 

Avoidance success decreased significantly in individuals exposed to these pharmaceuticals, thus 

potentially compromising survival and reproductive fitness. 
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Pharmaceutical Antibiotics

It is well known that antibiotic therapies can create antibiotic resistance in clinical settings as 

a result of repeated exposures of bacteria to individual antibiotics. The potential for antibiotics 

that have been released to the environment to enhance resistance patterns in receptor microbial 

communities is not well understood; however, recent research has focused attention on this 

topic. For example, D’Costa et al. (2011) discuss potential effects on resistance patterns of 

indigenous soil microbial communities exposed to antibiotics, and Duriez and Topp (2007) 

explore various manure management techniques at swine farms and their influence on antibiotic 

resistance patterns in the manure. Haack et al. (2012) showed that microbial community 

composition and structure can be affected by environmentally relevant exposures to an 

antibiotic (sulfamethoxazole). These investigators also noted that ecological function could be 

affected at concentrations two to three orders of magnitude smaller than those used in clinical 

applications and could promote antibiotic resistance through the selection of naturally resistant 

bacteria. In a related experiment, Underwood et al. (2011) showed that sulfamethoxazole 

exposures at environmentally relevant concentrations can affect the denitrifying activity 

of indigenous soil microbial populations in an aquifer that previously received wastewater 

effluents. LaPara et al. (2011) have identified tertiary treated municipal wastewater effluents 

to be significant point sources of antibiotic resistance genes to receiving surface waters in 

Minnesota. In another example of pharmaceutical antibiotic effects in the environment, Ebert 

et al. (2011) researched potential risks to photoautotrophic aquatic organisms (an alga, a 

cyanobacterium, and two macrophytes) resulting from exposure of fluroquinolone antibiotics 

(ciprofloxacin, enrofloxacin). The study identified risks (growth inhibition effects) to two of the 

test species at environmentally relevant concentrations of ciprofloxacin (risks were not identified 

at these concentrations for enrofloxacin). 

Conclusions

Existing data documenting the environmental occurrence of pharmaceuticals in the Chesapeake 

Bay watershed are limited so the extent and severity is identified as uncertain. However the 

limited data indicate a potentially wide extent of some pharmaceuticals for several reasons: 

1) the sources of pharmaceuticals, which include human and animal waste management 

operations such as municipal wastewater treatment effluents and animal agriculture, are widely 

distributed throughout the watershed; 2) the few studies that have been completed in the 

watershed have detected some pharmaceutical and related indicator chemicals in some streams, 

groundwater, and streambed-sediment; and 3) other studies conducted across the United States 
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and elsewhere consistently point to a wide extent of pharmaceuticals wherever source pathways 

to aquatic environments exist. The severity of pharmaceuticals as environmental contaminants 

is the subject of active research and very few studies have linked pharmaceutical exposures 

to individual or population-level mortality. However several studies in the U.S. and elsewhere 

have linked some of these chemicals to adverse sublethal effects such as predator avoidance 

behavior and incidence of intersex in a range of exposed organisms at environmentally relevant 

concentrations. In summary, sources of pharmaceuticals are fairly extensive in the watershed and 

studies on effects are increasing throughout the U.S. and therefore pharmaceuticals are important 

toxic contaminants of concern. However, data gaps and lack of complete understanding of 

potential adverse impacts continue to limit broader conclusions regarding extent and severity 

throughout the Chesapeake Bay watershed. 

2.7 Household and Personal Care Products

Abstract

Household and personal-care products (HPCPs) represent a wide range of organic chemicals 

that are used frequently in residences, workplaces, and other locations. Used in a variety of 

applications such as cosmetics, detergents, soaps, pest control, and food additives, these products 

are often formulated as a mixture of chemicals (including inorganic and organic active and 

inert ingredients). These contaminants can enter the environment in a number of ways but 

the predominant pathways are related to management and disposal of residential and human 

wastes such as landfills, on-site septic and municipal wastewater treatment plants. Findings 

included in the few local and regional occurrence datasets in the Chesapeake Bay watershed 

tend to mirror national studies of these compounds as potential environmental contaminants. 

For example, the compounds often detected in the environment include N,N-diethyltoluamide 

(insect repellant), surfactants (i.e. alkyl phenols), triclosan (an antimicrobial), and musks (tonalide, 

galaxolide). The environmental health implications of the presence of these contaminants in the 

environment are poorly understood; however, it is clear that: 1) these compounds often occur in 

complex mixtures; 2) the concentrations detected in environmental settings tend to be low as 

compared to the few existing health-based benchmarks; 3) some studies indicate that targeted 

compounds can produce estrogenic and other sublethal effects in exposed organisms (e.g. some 

surfactants, fragrances or musks, triclosan and triclocarban); and 4) due to the fact that sources 

of HPCPs are widely distributed throughout the built environment, it is reasonable to conclude 

that these compounds have the potential to occur extensively in receiving waters throughout the 

Chesapeake Bay watershed. Because of limited monitoring data and incomplete understandings 

of sublethal and other effects, extent and severity of HPCPs are identified as uncertain.
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Background

HPCPs include a wide range of inorganic as well as organic chemicals that are used frequently 

in residences, workplaces, and other locations. These products, such as cosmetics, detergents, 

soaps, and food additives, are typically formulated as a mixture of ingredients. For example, 

some soaps and toothpastes include antimicrobial compounds in their formulation (e.g., 

triclosan and triclocarban); some cosmetics include compounds commonly known as musks 

that provide fragrance (galaxolide and tonalide); detergents and cleaning products consist of 

several ingredients, including surfactants (nonylphenol ethoxylates, alkylphenol ethoxylates, 

and octylphenol ethoxylates); and other organic compounds (e.g., menthol) are used to enhance 

flavors in a range of products. 

Chemical constituents in HPCPs can enter the environment through a variety of pathways. 

Solid forms of these products are commonly placed in trash receptacles destined for landfills, 

whereas liquid forms are flushed in toilets or poured into sink drains. Therefore, most frequent 

source pathways to the environment are associated directly or indirectly with human-waste 

management activities such as municipal wastewater-treatment facility effluent discharges, septic 

tanks, and landfill leachates (Glassmeyer et al. 2008; Barnes et al. 2008). Chemical constituents 

associated with HPCPs can have a variety of sources and therefore are not always associated with 

household wastes (Phillips et al. 2012).

Water and Sediment 

These chemical constituents associated with HPCPs could derive from a variety of waste sources. 

Therefore, the analytes listed in this section are referred to as “indicators” of HPCPs when detected 

in the environment. Water quality standards generally do not exist for all chemicals that are 

indicators of HPCPs; therefore, these analytes are not commonly monitored in the Chesapeake 

Bay watershed. Several studies conducted throughout the United States and elsewhere, however, 

have provided monitoring data and associated ecological impacts (Vajda et al. 2011) and 

perspective on the few local data that are available. Most of the environmental data for these 

chemicals are for concentrations in water and some sediment samples. Tissue analyses are not 

currently available. 

Kolpin et al. (2002) were among the first to document the widespread occurrence of chemical 

indicators of HPCPs in streams in the United States. Glassmeyer et al. (2008) reviewed and 

summarized the available literature on the occurrence of pharmaceuticals and chemical 
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indicators of HPCPs in the environment worldwide and showed that 124 different compounds 

were detected in surface waters, 35 were detected in groundwater, and 4 were detected in bed 

sediments. Of the 126 compounds measured in at least one matrix, 18, including galaxolide and 

tonalide (which are used as fragrances in personal care products), were detected in more than 10 

different studies. 

Halden and Paull (2005) documented widespread occurrence of triclosan and triclocarban 

(antimicrobials) in water resources of the greater Baltimore area. Results of this study were 

compared to other locations throughout the United States and the authors concluded that these 

contaminants are widely distributed and frequently occurring. Elevated concentrations were not 

entirely due to incomplete removal of contaminants during wastewater treatment as maximum 

concentrations (triclocarban, 6.75 μg/L; and triclosan, 1.6 μg/L) in the Baltimore area were due to 

spills and leakage. 

Phelan and Miller (2010) reported results from base-flow, stormwater, and bed-sediment synoptic 

sampling of 23 sites in Rock Creek Park, Washington, DC. Of the indicators of HPCPs analyzed for, 

N,N-diethyltoluamide (insect repellant applied topically to clothing and skin), organophosphate, 

flame retardants, and galaxolide (musk) were detected at more than half the sites. Other 

indicators of HPCPs detected in water samples included alky phenol surfactants, triclosan 

(antimicrobial additive in soaps), and tonalide (musk). Compounds detected in bed-sediment 

samples include N,N-diethyltoluamide, galaxolide, and surfactants.

Alvarez et al. (2008) used passive samplers to assess the occurrence of a limited number of 

pharmaceuticals and other wastewater indicator compounds at nine locations in the Shenandoah 

River Basin and two in the James River Basin, Virginia. The objective of this study was to test the 

general hypothesis that declining fish health is linked to stream chemistry, but not to determine 

possible sources or source pathways. Detections of indicators of HPCPs included celestolide, 

galaxolide, phantolide, tonalide, and traseolide (musks used in a variety of cosmetics and 

other personal care products), N,N-diethyltoluamide, as well as several phosphate-based flame 

retardants used in a variety of consumer products such as clothes and furniture.

Barnes et al. (2008) reported datasets used in a national reconnaissance of pharmaceuticals 

and other organic wastewater compounds in untreated sources of drinking water in 25 states 

throughout the United States. Sites in the Chesapeake Bay watershed were located in Virginia, 
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Maryland, and Pennsylvania. Surfactants that are common in several types of detergents and 

cleaning products were detected at several sites. Other indicators of HPCPs detected included 

triclosan, menthol, and N,N- diethyltoluamide.

Reif et al. (2012) conducted reconnaissance sampling from 2006 to 2009 in selected locations 

across Pennsylvania to identify contaminants of emerging concern in (1) groundwater from 

wells used to supply livestock, (2) streamwater upstream and downstream from animal 

feeding operations, (3) streamwater upstream from and streamwater and streambed sediment 

downstream from municipal wastewater effluent discharges, (4) streamwater from sites within 5 

miles of drinking-water intakes, and (5) streamwater and streambed sediment where fish health 

assessments were conducted. Analytes included a range of indicators of HPCPs. 

Kolpin et al. (2013) analyzed discrete (grab samples) and integrated (passive samplers) water 

samples as well as bed sediment from smallmouth bass nesting sites in the Potomac River Basin. 

Chemical indicators of HPCPs detected in water samples include a range of surfactants, camphor, 

galaxolide, N,N-diethyltoluamide, tonalide, and triclosan.

Fish and Wildlife

Residues of HPCPs have yet to be reported in tissues of wildlife (amphibians, reptiles, birds, and 

mammals) in Chesapeake Bay. Ramirez et al. (2009) conducted a survey of five wastewatereffluent-dominated streams across the United States and analyzed fish tissue for a range of 

pharmaceuticals and HPCPs. Galaxolide and tonalide (musks) and triclosan were among the 

compounds detected.

Perfluorinated compounds are surface protectors and surfactants that are distributed worldwide. 

Perfluorooctanoic acid, an active ingredient of Scotch Guard®, was voluntarily removed from the 

marketplace in 2000 because it was increasingly being detected in the environment, wildlife, 

and people. Several perfluorinated compounds were detected in osprey eggs collected in the 

Chesapeake Bay watershed in 2000 and 2001, with upper extreme concentrations of 428 ppb 

wet weight for perfluorooctanesulfonate, 671 ppb for perfluorordecanesulfonate, 13.7 ppb for 

perfluorodecanoic acid, and 27.1 ppb for perfluoroundecanoic acid (Rattner et al. 2004). Recent 

controlled exposure studies and risk assessments in Northern bobwhite (Colinus virginianus) and 

mallards concluded that environmental exposure to perfluorooctanesulfonate does not pose a 

significant risk to avian populations (Newsted et al. 2005). 
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A series of alkylphenol and ethoxylate surfactants were also analyzed in 15 osprey eggs collected 

from the Chesapeake Bay watershed in 2000 and 2001. Nonylphenol was detected in all egg 

samples, with a maximum concentration of 16.7 ppb wet weight. Nonylphenol ethoxylates, 

octylphenol, and octylphenol ethoxylates were generally not detected, with the exception of 

octylphenol ethoxylate in 1 of 15 eggs at a concentration of 14.6 ppb. 

Lozano et al. (2012) documented seasonal variations in concentrations of alkylphenol and 

alkylphenol ethoxylate surfactants in tissue samples collected from largemouth bass living 

downstream from municipal wastewater-treatment plants in Chicago, IL. In comparison, the 

tissue concentrations of other persistent organic compounds such as PCBs, polybrominated 

diphenyl ethers (PBDEs), and some pesticides such as DDT, did not vary over time. Although 

these analyses were not done in the Chesapeake Bay watershed, the contaminants associated 

with wastewater (e.g., surfactants) are similar. 

The toxicological effects of HPCPs in the environment are an area of active research. Some 

compounds, such as the antimicrobials triclosan and triclocarban, have been shown to have 

sublethal effects at environmentally relevant concentrations ranging from stimulation of embryo 

production in freshwater snails (Potamopyrgus antipodaru) (Guidice and Young 2009) to disruption 

of thyroid homeostasis (Dann and Hontela 2010). Dann and Hontela (2010) also point out that 

mammalian toxicity studies of triclosan have shown that it is not acutely toxic, carcinogenic, 

mutagenic, or a developmental toxicant; however, they also note that some aquatic species such 

as algae, invertebrates, and some fish are sensitive to triclosan. Bedoux et al. (2012) reviewed 

the literature on triclosan toxicity and environmental occurrence and showed that benchmarks 

such as EC50 generally were much lower (μg/L or lower) in plants than in animals, with toxicity 

to animals well above most environmentally relevant concentrations. On the other hand, some 

indigenous soil bacteria and aquatic algal species did exhibit sublethal and other effects (e.g., 

inhibition of soil microbial respiration) at environmentally relevant concentrations. Evidence of 

the biological effects of HPCPs in the environment can be found for other contaminants as well. 

For example, surfactants have been identified as contributing to the estrogenicity of wastewater 

effluent and endocrine disruption in Boulder, Colorado (Vajda et al. 2011) and Lozano et al. (2012) 

showed that APE surfactants can elevate plasma vitellogenin in male largemouth bass living in 

streams receiving wastewater effluents. Breitholtz et al (2003) concluded that synthetic musks are 

not likely to adversely impact copepods at environmentally relevant concentrations. 
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However, Carlsson and Norrgren (2004) studied the toxicity of synthetic musks to early life stages 

of zebrafish (Danio rerio) and concluded that some musks can have adverse impacts on wild fish 

at environmentally relevant concentrations.

Conclusions

HPCPs represent a group of contaminants with many potential uses and a diverse array of 

chemicals. Many studies throughout the world have shown these contaminants to be associated 

with a wide range of sources such as municipal wastewater effluents, on-site septic tanks, and 

other waste management operations (e.g. landfills). Therefore, although the existing occurrence 

data are limited in the watershed, the sources such as wastewater effluents are widespread 

indicating a potential for some of the commonly used chemicals within this group to occur in 

many locations throughout the watershed. Significant gaps exist in understanding the range of 

potential adverse ecological effects of HPCPs, their degradates and mixtures, at environmentally 

relevant concentrations thereby limiting broader conclusions about their severity in the 

Chesapeake Bay watershed.

2.8 Polybrominated Diphenyl Ether Flame Retardants

Abstract

Polybrominated diphenyl ethers (PBDEs) have been used as flame retardants in a range of 

consumer products since the 1970s. As a contaminant group, PBDEs include more than 200 

potential congeners, making this one of the most extensive groups in this report in terms of 

total number of individual chemicals. Because PBDE congeners are highly hydrophobic, they are 

more likely to be found in sediments and tissue than in the water column. PBDEs are known to 

bioaccumulate; therefore, environmental monitoring for PBDEs has focused on animal tissue (e.g., 

fish, shellfish, and a range of terrestrial animals) in the Chesapeake Bay watershed and around 

the world, including remote regions of the Arctic. Because of their documented persistence and 

tendency to bioaccumulate, along with concerns about potential toxicity, certain penta- and 

deca-BDE formulations were phased out in 2004 and deca-BDE formulations will be phased 

out by the end of 2013. Also known to exist in coastal marine sediments globally, PBDEs are 

well documented in terms of their extent elsewhere. and therefore have the potential to be 

widespread in Chesapeake Bay. On the other hand, the potential severity of toxicological and 

ecological effects of PBDEs as environmental contaminants are less well understood. Currently 

there are no fish consumption advisories for PBDEs in the Chesapeake Bay watershed; however, 

adverse sublethal effects at environmentally relevant concentrations have been documented 
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in a limited number of test organisms. Because of limited monitoring data and incomplete 

understandings of sublethal and other effects, extent and severity of PBDEs are identified as 

uncertain.

Background

Polybrominated diphenyl ether (PBDE) flame retardants have been commonly used as fire 

retardants in textiles, electronics, polymers, and other materials since the 1970s. Although the 

deca-BDE mixture has been produced in the greatest volumes, the demand for penta-BDE in 

North America has exceeded that in other markets by an order of magnitude (Hale et al. 2006). 

Pathways for PBDEs to enter the environment are poorly understood and quantified. Because 

PBDEs are chemically incorporated into consumer products, the most likely environmental 

pathways include air emissions from manufacturing and product usage. Hale et al. (2002) 

suggest that these compounds enter the environment as polyurethane foams, disintegrate, and 

release congeners in the resulting fragments. 

Because PBDEs are hydrophobic and lipophilic, the majority of environmental occurrence studies 

for PBDEs are focused on animal tissue and soils or bed sediment. Hale et al. (2002) found 

elevated concentrations of PBDEs in soils, bed sediment, and bluegill tissues downwind and 

downstream from a polyurethane foam manufacturing plant in the Roanoke River watershed 

(outside the Chesapeake Bay watershed). Due to their persistence, bioaccumulation and 

potential toxicity, USEPA and manufacturers reached agreement to phase out the use of certain 

penta- and octa-BDE formulations by 2004, and more recently agreed to phase out the use of the 

deca-BDE formulation by the end of 2013 (USEPA 2012d). 

Water and Sediment

Because PBDEs have low solubility and there are no regulatory standards for PBDEs, no watercolumn data for these compounds were identified in the literature or found in any of the data 

searched for this report. 

Kimbrough et al. (2009) measured concentrations of PBDEs (mono- through hepta- congeners) 

in paired sediment and oyster tissues (see below) collected at the NOAA nationwide Mussel 

Watch stations from 2004 to 2007. The Chesapeake Bay was placed in the lowest of three clusters 

(low, medium, and high) for paired sediment and tissue concentration data. In the six Maryland 

stations in the Bay, four had no detectable PBDE concentrations in sediments; the other two were 
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designated as medium concentrations of 0.1 and 0.3 ppb dry weight. In the four Virginia samples, 

two were non-detects one had 0.1 ppb, and one had 0.4 ppb (both designated as medium 

concentrations). 

Fish and Shellfish 

According to the ATSDR (2004), the main source of human exposure to PBDEs may be through 

foods with high fat content, such as fatty fish. Some lower brominated PBDEs have been detected 

in air samples, indicating that people can also be exposed by inhalation. The ATSDR (2004) public 

health statement notes that rats and mice exhibited thyroid effects after consuming food with 

moderate amounts of lower brominated PBDEs for short periods. Rats and mice that ate smaller 

amounts over longer time periods (weeks, months) had liver and thyroid effects. The statement 

speculates that the thyroid effects are specific to the test animals and may not be relevant for 

humans. Understanding of the carcinogenicity of PBDEs is incomplete although USEPA lists 

decaBDE as a possible human carcinogen. Birth defects have not currently been linked to PBDE 

exposures but research is on reproductive effects is ongoing. 

There are no fish tissue advisories for PBDEs in the Chesapeake Bay watershed. PBDE 

concentrations were included as part of monitoring conducted to update fish tissue advisories for 

District of Columbia (Pinkney 2009). In 2007, samples of the following species were collected from 

the Potomac and Anacostia Rivers: sunfish, largemouth bass, carp, blue catfish, channel catfish, 

and American eel. Total PBDE concentrations were compared with the Virginia Department of 

Health (VDH) Guidance Value of 5 ppm (R. Tripathi, VDH, personal communication) because no 

USEPA fish-tissue screening value exists for PBDEs. The maximum concentration was 0.136 ppm in 

a channel catfish sample. No concentration in any of the other samples exceeded 0.1 ppm.

Klosterhaus and Baker (2010) examined the bioavailability of decabromdiphenyl ether (BDE 209) 

to the marine polychaete worm, Nereis virens. Bioaccumulation of this compound was minimal 

but uptake of two degradation products (BDE 207 and 208) did occur. Klosterhaus et al. (2011) 

reported much higher bioaccumulation of penta-BDE, with biota-sediment accumulation factors 

as high as 0.6 compared with of 0.0003 for BDE 209. This is consistent with studies showing tissue 

residues of tetra- and penta-BDEs in wildlife.

In their review of the literature on the toxicity of PBDEs, Wenning et al. (2011) stated that 

PBDEs are expected to be toxic to aquatic organism but dose-response relations have not been 
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developed. Recent studies of single PBDE compounds in the laboratory demonstrate a range of 

sublethal effects. For example, Kuiper et al. (2008) exposed zebrafish (Danio rerio) to a purified 

pentabromodiphenylether (congener BDE-71) in water in a 30-day exposure that focused on egg 

production and fertilization. Eggs were hatched and exposure continued until 45 days posthatch. The authors also measured whole-body BDE-71 concentrations in adults and juveniles. 

They found that, although larval survival was significantly reduced, the adult tissue concentration 

associated with the exposure (126 ppm wet weight) was more than 55 times the highest reported 

environmental concentrations in wild fish (burbot, Lota lota in a Norwegian lake at 2.27 ppm wet 

weight). In contrast, Chou et al. (2010) reported adverse effects on swimming speed and level 

of activity in juvenile zebrafish exposed to the pentabromodiphenylether BDE-47 from day 21 

through day 54 post-hatch. They stated that these effects occurred at concentrations approaching 

those reported in environmental samples of wild fish.

As stated above, Kimbrough et al. (2009) measured concentrations of PBDEs (mono- through 

hepta- congeners) in paired sediment and oyster tissues collected at NOAA’s nationwide Mussel 

Watch stations from 2004 to 2007. The six Maryland stations were designated as having medium 

concentrations in oyster tissue (7.1–24.1 ppb total PBDEs). All five Virginia sites had medium 

concentrations ranging from 8.5 to 14.8 ppb. Kimbrough et al. (2009) concluded that the toxicity 

and ecosystem effects of PBDEs on marine biota have not been well studied. 

Wildlife

In osprey eggs collected in 2000 and 2001, BDE congeners 47, 99, 100, 153, and 154 were 

detected in all samples from Chesapeake Bay urbanized areas and the South, West, and Rhode 

Rivers reference sites near Annapolis, Maryland (Rattner et al. 2004). Total PBDE concentrations 

ranged up to 928 ppb wet weight (some of the greatest concentrations reported in bird eggs to 

date), and approach the lowest-adverse-effects-level of 1,800 ppb eggs for pipping and hatching 

success derived in American kestrels (McKernan et al. 2009). Tetra- and penta-BDE congeners 

typically dominate eggs of piscivorous birds. In a collection of addled peregrine falcon eggs from 

13 nests in the Chesapeake Bay watershed (Potter et al. 2009), total PBDE concentrations ranged 

from 32.8 to 354 ppb wet weight. The deca-BDE congener 209 was detected at concentrations 

ranging up to 48.2 ppb, and levels of this congener were found to be positively correlated with 

human population density in proximity to the nests (Potter et al. 2009). Polybrominated biphenyl 

congener 153 was also quantified in many of the peregrine eggs, ranging from 5.88 to 51.8 ppb 

wet weight, although toxicity thresholds in avian eggs are unknown. In 2010, six common tern 
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eggs collected from Poplar Island were analyzed for PBDEs. These eggs contained low levels of 

BDE congeners (total PBDE ranged from 10.2 to 51.0 ppb wet weight), making them ideal for their 

intended use in a controlled exposure penta-BDE egg injection (Rattner et al. 2011).

Congeners of PBDEs detected in penta- and octa-BDE formulations bioaccumulate and 

biomagnify in food chains (de Wit 2002), and detailed investigations in captive and free-ranging 

birds have documented developmental, endocrine, immunologic, and reproductive effects at 

environmentally relevant concentrations (Chen and Hale 2010). 

Conclusions

Limited data suggests that PBDEs occur in the tissue of aquatic and terrestrial organisms and 

in sediment; however, the amount of data is not sufficient to determine extent of occurrence 

in the Bay. It is notable that recent research on the adverse effects of PBDEs is focusing on 

specific congeners and results are varied. Research on sublethal effects would inform broader 

conclusions as to their severity as potential environmental contaminants in the Chesapeake Bay 

watershed. Because of limited monitoring data and incomplete understandings of sublethal and 

other effects, extent and severity of PBDEs are identified as uncertain.

2.9 Biogenic Hormones

Abstract

By definition, biogenic hormones are created, biologically processed, and excreted by humans 

and other organisms. Therefore, there is potential widespread extent of occurrence of some 

biogenic hormones in the environment. Recent monitoring studies have focused on natural 

and synthetic estrogens as well as several other chemicals that are known or suspected to 

interfere with normal estrogen receptor functioning in a range of organisms. Monitoring results 

have shown that municipal wastewater effluents as well as animal agricultural activities can be 

major pathways for biogenic hormones to reach the environment. Other studies have shown 

that exposures to these environmental hormones cause intersex and other abnormalities such 

as immune suppression in fish and other species. Although these compounds do not have 

immediate lethal effects at environmentally relevant concentrations, they may have more subtle 

effects on developmental and maturational processes and may diminish individual fitness 

and reproductive life span. These key sublethal effects are typically difficult to detect in wild 

populations and present an insidious challenge to research on the health of fish and wildlife. 

Currently (2012), the weight of evidence indicates that environmental exposures to biogenic 
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hormones can cause adverse effects to some receptor organisms. The occurrence data on 

biogenic hormones in water and sediment is presently limited to a few focused studies and is 

identified as uncertain. Conclusions about the severity of biogenic hormones in the Chesapeake 

Bay watershed are limited by the available data and research. Because of limited monitoring data 

and incomplete understandings of sublethal and other effects, extent and severity of biogenic 

hormones are identified as uncertain.

Background

The presence of natural hormones in the environment, a result of animal and plant waste 

excretion and other natural biological processes, can be anthropogenically enhanced by various 

waste-management practices. As a compound group, natural hormones are not commonly 

monitored by regulatory agencies as contaminants; however, the combination of potentially 

large quantities of these compounds reaching the environment as a result of waste-management 

activities and the fact that they are biologically active at low concentrations has increased 

research attention on these compounds as potential environmental contaminants.

All multicellular organisms including plants and animals produce hormones. Hormones are 

required for a variety of biological functions and include, for example, estrogen and androgen, 

which are critical for normal sexual reproduction in vertebrate animals. Other examples of 

animal hormones include tyrosine-based hormones associated with thyroid functioning and 

corticosteroids produced in the adrenal cortex and associated with a range of stress and immune 

responses and metabolic processes. Phytoestrogens are produced by plants and, although they 

are weakly estrogenic compared to animal estrogens, they are produced in abundance, and their 

environmental pathways also are affected by waste-management activities. 

This section of the report focuses on naturally produced hormones because of current related 

research on fish health and intersex in the Chesapeake Bay watershed (Ciparis et al. 2011; Blazer 

et al. 2012; Kolpin et al. 2013) and elsewhere in the world including studies that have linked 

estrogenic contaminants and intersex in fish (Vajda et al. 2011). Current research is focused on 

the naturally occurring steroidal hormones as well as other environmental contaminants that bind 

to or otherwise affect estrogen and androgen receptors; however, there is a growing list of other 

chemical compounds (e.g. some personal care products, pharmaceuticals, etc.) that are known or 

suspected to interact with hormone receptor signaling pathways. These other compounds are not 

considered in this section of the report, but are discussed in other sections.
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Because these compounds are key components of all animal and plant life, their environmental 

sources are ubiquitous. Some management activities associated with human, animal, and plant 

wastes and associated biological material can be key point and(or) nonpoint source pathways 

for hormones in the environment. Municipal wastewater treatment plants are centralized 

locations where human wastes (and associated hormones) are treated before being discharged 

to the environment. These treatment systems are not specifically designed to remove hormones 

and consequently some municipal wastewater effluents are potential point sources of natural 

hormones to the environment. Consequently successful removal (or biochemical transformation) 

of parent hormone compounds as they pass through existing wastewater treatment plants can 

be highly variable (Chimchirian et al. 2007). Effluent from WWTPs can be a significant source 

of natural estrogens from human excretion to surface waters (Lagana et al. 2004; Muller et al. 

2008; Petrovic et al. 2002; Ying et al. 2008; Phillips et al. 2012). Finally, because sewersheds 

can encompass large land areas and various sources, the treatment plants can become central 

locations where hormone-laden wastes accumulate and potentially become concentrated before 

release to the environment.

Other sources include land uses and related operations where animal manure is created and 

accumulated (e.g. lagoons at livestock feeding operations) and inadvertently (or intentionally) 

released to the environment as point sources. Rangeland or pasture grazing of livestock can be 

another source of hormones to the environment especially where livestock have direct access 

to streams or other water bodies (Kolodziej and Sedlak 2007). Other agricultural activities 

such as where animal manures are spread on cropland as a source of soil fertility, biosolid uses, 

and compost operations accumulate animal and plant wastes and eventually release them to 

the environment in various forms and conditions of decomposition. The latter are important 

nonpoint sources of hormones to surface and groundwater during runoff and leaching events 

(Finlay-Moore et al. 2000; Kjaer et al. 2007; Matthiessen et al. 2006; Shore et al. 1995). 

Metabolic processes associated with plants and animals produce a range of parent and 

conjugated forms of hormones. Therefore, hormones excreted from animals can be in the original 

parent form as well as a range of conjugated forms such as glucoronides and sulfur compounds, 

which in turn are subject to further transformation processes. Where environmental conditions 

are amenable, the conjugated forms can be cleaved to the parent compound and other processes 

can produce the conjugated forms again. Several studies have documented concentrations 

of parent steroids as well as their glucoronide and sulfate conjugated metabolites in livestock 
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manures and municipal biosolids destined for land applications (Hanselman et al. 2003; Andaluri 

et al. 2012). The glucoronide metabolites have been shown to deconjugate back to the parent 

steroid in the environment (Ternes et al. 1999; Panter et al. 1999). In contrast, current evidence 

on sulfate conjugates indicates they are much more resistant to transformation (Johnson and 

Williams 2004; D'Ascenzo et al. 2003). Hutchins et al. (2007) analyzed samples from several 

different lagoons at confined animal-feeding operations (CAFOs) and concluded that estrogen 

conjugates contribute to the overall estrogen load. Therefore, fate and transport studies of 

natural plant and animal hormones must account for the parent as well as the conjugated 

forms, especially when their role as hormonally active environmental contaminants is being 

investigated. 

Hanselman et al. (2003) completed a literature review to assess the current state of science 

regarding estrogen physicochemical properties, livestock excretion, and the fate of manureborne estrogens in the environment. Unconjugated steroidal estrogens have low solubility in 

water (0.8-13.3 (mg/L) and are moderately hydrophobic (log Kow 2.6-4.0). Cattle excrete mostly 

17α-estradiol, 17β-estradiol, estrone, and respective sulfated and glucuronidated counterparts, 

whereas swine and poultry excrete mostly 17β-estradiol, estrone, estriol, and respective sulfated 

and glucuronidated counterparts. 

Khanal et al. (2006) reviewed the literature on the fate, transport, and biodegradation of natural 

estrogens in the environment. Although the environmental fate of estrogens is not clearly 

known, Hanselman et al. (2003) summarized the results of laboratory-based studies that found 

the biological activity of these compounds is greatly reduced or eliminated within several hours 

to days as a result of degradation and sorption. Bradley et al. (2009) studied the biodegradation 

of 17β-estradiol, estrone, and testosterone in three streams in the United States that receive 

wastewater-effluent effluents. The results showed that biodegradation of these hormones 

in conjunction with sorption to bed sediments can be mechanisms for the attenuation of 

hormonal endocrine disruptors in effluent-affected streams. Gray and Sedlak (2005) showed that 

constructed wetlands function to remove or transform estrogenic contaminants. Approximately 

36% of the 17β-estradiol was attenuated, with the most likely cause being sorption to 

hydrophobic surfaces in the wetland coupled with biotransformation. However, the attenuation 

processes in these studies were still not sufficient to completely mineralize or sequester these 

compounds. Other studies (Khanal et al. 2006) have demonstrated that estrogens are sufficiently 

mobile and persistent to affect surface water and groundwater quality and, therefore, aquatic 
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organisms can be exposed to these hormones even in locations that are removed from their 

sources or pathways to the environment.

Khanal et al. (2006) found that conventional wastewater treatment is efficient in the removal 

of 17β-estradiol, but estrone removal is relatively poor. The removal occurs mainly through 

sorption by sludge and subsequent biodegradation. Barber et al. (2012) showed that estrogenic 

compounds were removed from treated wastewater by upgrading a wastewater-treatment 

plant from a trickling-filter/solids-contact process to an activated-sludge process. The removal 

was in large part, the result of increases in hydraulic retention time and solids retention time. 

Before the upgrade, 17β-estradiol was partially removed by the trickling filter, whereas estrone 

concentrations actually increased in the effluent, likely as a result of biologically mediated 

oxidation of 17β-estradiol. Concentrations of estriol, 17β-estradiol, and estrone in treated effluent 

decreased to below detection levels after the upgrade. 

Studies and monitoring data on steroidal hormones in the Chesapeake Bay watershed are limited. 

A few example studies conducted by Federal agencies and academia are cited below; however, 

available information is insufficient to warrant individual subsections on water, sediment, and 

tissue. 

Water and Sediment

Many studies have documented the presence of steroidal hormones in aquatic environments 

(Kolpin et al. 2002). Most of these studies have focused on dissolved constituents in the water 

column, with fewer studies of those in bed sediment. Kinney et al. (2006) analyzed biosolids 

produced from wastewater-treatment-plant sources and detected several steroidal hormones 

including 3-β-coprostanol, cholesterol, β-sitosterol, and stigmastanol at high concentrations 

relative to the concentrations of other organic compounds detected. These commercially 

available biosolids are commonly applied to land surfaces as soil amendments and are therefore 

potential source pathways for steroidal hormones to the environment. Data on phytoestrogen 

occurrence in aquatic environments is much more limited; however, Kolpin et al. (2010) analyzed 

water from 15 streams draining agricultural land in Iowa for a range of phytoestrogens. Target 

compounds that were frequently detected included formononetin (80%), equol (45%), daidzein 

(32%), biochanin A (23%), and genistein (11%).
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Reif et al. (2012) analyzed 270 streamwater samples collected near drinking-water intakes in 

Pennsylvania for a range of biogenic hormones. Six of the 17 hormones and 2 of the animal 

sterol compounds were detected at least once. The most frequently detected hormones were 

estrone (18% of samples; maximum concentration 3.1 ng/L), cis-androsterone (5% of samples; 

maximum concentration 6.2 ng/L), and 4-androstene-3,17-dione (3% of samples; maximum 

concentration 1.8 ng/L). All other hormones were detected in fewer than 1 percent of samples. 

Only 1.5% of all 5,130 analyses resulted in hormone detections. Dorabawila and Gupta (2005) 

documented the presence of estradiol (17β-estradiol) in water samples from the Wicomico, 

Manokin, and Pocomoke Rivers with highest concentrations downstream from sewage-treatment 

plants. 

Ciparis et al. (2012) sampled tributaries in the Shenandoah River watershed during three different 

seasons and analyzed the water by using an estrogen assay. The bioluminescent yeast estrogen 

screen (BLYES) indicates the presence of compounds that can bind to the estrogen receptor. 

In this way, BLYES is an indicator of “total estrogenicity” of the water and includes activities 

associated with all natural and synthetic estrogens or other chemicals that are estrogen agonists 

or antagonists. Consequently results of these assays cannot entirely be attributed to biogenic 

hormones; however, for simplicity the total concentration of estrogenic compounds in the water 

samples is reported relative to 17β-estradiol (E2) as E2 equivalents (E2Eq). Concentrations in 18 

samples from 10 sites were >1 ng/L E2Eq, and animal feeding operation (AFO) densities in the 

watersheds of most of these sites were >1 per 1,000 acres. An E2Eq > 1 ng/L was the predicted 

no-effect concentration of total estrogens on fish reproduction. This study documented 

statistically significant (R2=0.39-0.75) relations between watershed densities of AFOs and instream concentrations of E2Eq. The range in estrogenic activity observed during this study was 

similar to estrogen concentrations and total activity measured in other studies across the United 

States, including (1) streams with concentrated sources, such as streams draining fields that 

receive wastes from poultry AFOs; (2) streams adjacent to pastures receiving waste from dairy 

and beef AFOs; and (3) rangeland streams directly accessed by grazing cattle. A similar range in 

estrogenic activity has been measured in treated WWTP effluent (Salste et al. 2007) and in rivers 

receiving effluent from WWTPs (Vermeirssen et al. 2005).

Arikan et al. (2008) analyzed 26 water samples from seven river stations and 56 water samples 

from 15 subwatershed stations in the Choptank River Basin for several biogenic hormones. Estriol 

(12 ng/L) and estrone (13 ng/L) were each detected once in subwatersheds, and progesterone 
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was detected in three different subwatersheds (12, 12, and 14 ng/L). The synthetic estrogen 

17α-ethinylestradiol was also detected in one subwatershed water sample. Testosterone (16 ng/L) 

was detected once at a river station. One other hormone (17β-estradiol) was analyzed for but not 

detected in any water sample. 

As part of a national AFO study, the USGS has monitored runoff from streams and ditches draining 

a pastured cattle operation in the Rappahannock River Basin. Samples were collected during a 

range of high and low flows by using both grab samples and passive samplers. Preliminary results 

indicate detections of steroidal hormones in water and bed sediment, including cholesterol, 

cis-androsterone, coprostanol, estrone, epitestosterone, testosterone, 4-adrostene-3, 17-dione, 

17α-estradiol, and 17β-estradiol (D.W. Kolpin, USGS, personal communication).

In a continuing investigation regarding the potential connections between contaminants and 

fish health and intersex in the Chesapeake Bay watershed, the USGS collected samples of water 

and bed sediment at six active smallmouth bass nesting stream sites and one control site during 

spawning (Kolpin et al. 2013). Although hormone analyses in this study have been limited, 

the steroidal hormones detected in water and sediments include 17α-estradiol, 17β-estradiol, 

cholesterol, estrone, sitosterol, and stigmastanol. In addition, estrogenicity of sediments 

collected at nest sites was positively correlated with incidence of fish intersex.

Fish 

Although natural endogenous steroidal hormones are present in all animals, hormone 

concentrations in tissue such as blood plasma are potentially useful indicators or markers of 

exposures to environmental or exogenous sources of hormones. Iwanowicz et al. (2009b) 

analyzed blood plasma of smallmouth bass collected from two tributaries in the Chesapeake Bay 

watershed for an estrogen (17-β estradiol) and an androgen (testosterone). Fish upstream and 

downstream from municipal wastewater effluents were compared. The general increase in blood 

plasma hormone concentrations in fish downstream from wastewater sources indicates likely 

exposures to these or other hormonally active compounds in the wastewater effluents.

Citing weight of evidence for endocrine disruption in fish, the Environment Agency of 

England and Wales recently focused on risk-management strategies for natural (and all other) 

estrogenically active effluents that discharge to the aquatic environment (Gross-Sorokin et al. 

2006). To date (2012), most studies of environmental hormones have focused on estrogenicity 
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and related effects. Khanal et al. (2006) cited estrogenicity monitoring studies in more than 30 

countries and concluded that natural steroidal estrogens such as estrone, 17β-estradiol, estriol, 

and 17α-estradiol are potent endocrine disrupters found in the environment. Phytoestrogens 

are structurally similar to 17β-estradiol (estrogen in vertebrate animals) and their interaction 

with vertebrate estrogen receptors is a topic of active research in human and ecological health 

sciences. Currently (2012), research on other natural hormones as potential environmental 

contaminants is limited.

Yonkos et al. (2010) exposed fathead minnows (Pimephales promelas), sheeepshead minnows 

(Cyprinodon variegatus), and mummichogs (Fundulus heteroclitus) to solutions leached from 

poultry litter. The minnows exhibited intersex conditions whereas the mummichogs were 

unresponsive in all trials. Again, the potential mixture of hormonally active chemicals in the litter 

likely includes biogenic hormones as well as other endocrine disruptors; however, this study did 

not test for individual chemicals. In a positive controlled exposure experiment, 17β-estradiol 

(biogenic estrogen) had no effect on the mummichogs but did have effects on the minnows 

(Yonkos et al. 2010). 

Recent research has linked environmental exposures of estrogenic compounds to fish intersex 

in Boulder Creek, Colorado, and the Potomac River Basin, Virginia (Vajda et al. 2008; Blazer et al. 

2011; Kolpin et al. 2013). Because the sources of biogenic hormones in human and animal wastes 

are consistently associated with other estrogenic chemicals, these intersex studies have not linked 

specific effects to individual naturally occurring animal or plant hormones. For example, among 

the targeted chemicals analyzed by Vajda et al. (2008) that are known or suspected endocrine 

disruptors, 17β-estradiol, estrone, estriol, and 17α-ethynylestradiol as well as estrogenic 

alkylphenols and bisphenol A were identified and their mixtures were linked to intersex 

conditions in fish. Vajda et al. (2008) also showed that as separate groups the steroidal estrogens 

and the nonsteroidal estrogens were both in concentrations high enough to disrupt normal 

endocrine functioning. However, VanDenBelt et al. (2004) compared the estrogenic potencies 

of 17β-estradiol, estrone, 17α ethinyl estradiol, and nonylphenol (surfactant) and showed that 

17α ethinyl estradiol can be as estrogenic as, or many times more estrogenic than 17β-estradiol 

(estrogenic potencies are commonly reported relative to 17β-estradiol), but nonylphenol is likely 

orders of magnitude less potent. As stated above, Kolpin et al. (2013) analyzed bed sediment 

at smallmouth bass nest sites in the Potomac River watershed and found a positive correlation 

between total biogenic estrogens (17β-estradiol, estrone, sistosterol, stigmastanol, and 
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progesterone) in the sediments and incidence of intersex in fish using those nests. Therefore, the 

biogenic estrogens are likely critical components of total estrogenicity in terms of potency and 

therefore must be considered key environmental contaminants.

An additional effect of hormones is their role as potential promoters of liver tumors in fish. Nunez 

et al. (1989) first demonstrated that 17β-estradiol promotes hepatocellular carcinoma in fish. 

Cooke and Hinton (1999) reported greater liver tumor prevalence in females of six fish species 

and theorized that endogenous estrogens serve as tumor promoters. Although the effect of 

estrogens on liver tumors in the Chesapeake Bay watershed is unclear, Pinkney et al. (2011) found 

a similar higher prevalence of liver tumors in females versus males and incorporated sex as a 

covariate in a logistic regression analysis of prevalence across sampling locations.

Wildlife

A primary mode of exposure to exogenous steroids or environmental chemicals in birds is through 

maternal deposition (Adkins-Regan et al. 1995; Ottinger et al. 2005), which is significant across 

a range of endocrine disruptors and contaminants, with the distribution of the toxicant in the 

egg dependent on the chemical characteristics of that compound (Lin et al. 2004; Ottinger et al. 

2005). This potential route of exposure is supported by a study in which Japanese quail hens were 

given estradiol implants (Adkins-Regan et al. 1995). Assay of plasma samples showed increased 

circulating estradiol levels, with either daily estradiol injections or with a silastic implant containing 

crystalline estradiol. Eggs produced by treated females contained significantly greater estradiol 

concentrations in the yolks than those produced by control females; therefore, these data provide 

evidence for maternal transfer of steroid hormones to the offspring through the yolk (AdkinsRegan et al. 1995). Similarly, other lipophilic compounds, including the soy phytoestrogens, also 

transfer from the hen into the egg and. more specifically, the yolk (Lin et al. 2004). 

Embryonic exposure to exogenous steroid hormones alters sexual differentiation in birds and 

results in reproductive impairment in adults (Adkins-Regan et al. 1990, 1995; Halldin et al. 1999; 

Ottinger and vom Saal 2002; Ottinger et al. 2005). Embryonic exposure to estradiol or androgen 

greatly affects sexual differentiation, whereas adult exposure is ineffective (Adkins-Regan et al. 

1990; Ottinger and Abdelnabi 1997). Similar results have been observed in bobwhite quail and 

also in mammals, indicating that this effect may be common across species and even phyla (Lien 

et al. 1987). 
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Conclusions

Although only few studies have been conducted on the occurrence of biogenic hormones in 

the watershed, it is clear that these naturally occurring compounds are widespread. Previous 

studies have shown that wastewater management operations such as municipal sewer 

systems can result in point sources of biogenic hormones that are reflective of the combined 

inputs within the sewershed. Other studies have shown that upgrades in treatment plants 

can lower the estrogenicity concentrations of the effluents. Animal agricultural practices, 

particularly those associated with manure management and others where livestock have direct 

access to streams are also known sources of these contaminants. In addition, although these 

contaminants are naturally occurring they are likely to be elevated in concentrations above 

background downstream from these pathways to the environment. Hormones are known 

to be physiologically active at very low levels and have been shown to cause adverse effects 

on receptor organisms at environmentally relevant concentrations. There are many potential 

sources and pathways of biogenic hormones to the environment in the Chesapeake Bay 

watershed therefore they likely are widely distributed in water and sediment. Current research 

is addressing how these contaminants may be impacting aquatic environments but at this 

time there is insufficient information to assess the overall severity of biogenic hormones in the 

watershed. 

2.10 Metals and Metalloids

Abstract

Parsing the effects of metals in the environment is a complex process. Metals occur naturally 

in all media and are typically present in the environment as mixtures. Chemical and biological 

processes can alter metals valence states and can convert them between organic and inorganic 

forms. These factors, along with different physical and chemical properties of each metal, makes 

environmental risk assessment challenging. Although the concentrations of trace metals have 

declined slowly since the 1980s, elevated concentrations remain in some regions of the Bay. The 

most prevalent cause of impairment is the presence of mercury in fish tissue at concentrations 

in excess of State guidelines, affecting more than 600 river miles and approximately 20,000 

impoundment acres in the Chesapeake Bay watershed. In the water column, other metals 

are present at concentrations in excess of State standards for the protection of aquatic life. 

Specifically, both Maryland and West Virginia have identified impairments for aluminum and iron 

in water, largely limited to each jurisdiction’s portion of the North Branch of the Potomac River. 

Maryland also included manganese as a cause of impairment in the North Branch of the Potomac 
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River. In addition, Maryland’s Bodkin Creek is identified as impaired for copper. Pennsylvania has 

completed more than 70 TMDLs for metals impairments, covering several hundred miles within 

the bay watershed. The great majority of Pennsylvania’s metal impairments are for aluminum, 

iron, and manganese originating from abandoned mine discharges in the coal areas of the upper 

Susquehanna watershed. Maryland listed additional impairments for chromium, lead and zinc in 

sediment (Baltimore Harbor, Bear Creek and Curtis Bay Creek). These impairments were minimal in 

extent; zinc was the most widespread contaminant, impacting approximately seven square miles 

of estuarine waters. Based on the spatial extent of the listed impairments, the extent and severity 

of mercury contamination within the Bay watershed is considered to be widespread, whereas 

contamination with other metals (i.e., aluminum, chromium, iron, lead, manganese and zinc) is 

more localized. 

Background

Metals are ubiquitous in the environment as constituents of both rock and sediment, and 

concentrations vary across geographic regions. They may be delivered to aquatic environments 

through natural erosion and weathering but also from anthropogenic sources. When 

concentrations in aquatic environments exceed natural background levels as a result of human 

activities, metals are considered to be contaminants. The term “metals” as used in this document 

refers to true metals, metalloids and organometals. 

Risk from metals contamination in aquatic environments is complex because of the unique nature 

and behavior of elements. For example, because metals occur naturally in all environmental 

media, they typically are present in the environment in mixtures. Some metals are nutrients that 

are essential for life and maintaining the health of humans, animals, plants, and microorganisms. 

Unlike organic contaminants, metals are not degraded or destroyed by chemical or biological 

processes, but instead those processes can transform metals from one species to another (valence 

states) and can convert them between organic and inorganic forms. These factors, along with 

differing characteristics for each metal, create highly variable biological and ecological effects in 

the environment, making ecological risk assessment uniquely challenging. Extended exploration 

of these topics is beyond the scope of this report. Moreover, this chapter is not intended to be an 

exhaustive review of all metal contaminants; its focus is on those metals that have been found 

within the Chesapeake Bay watershed.
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According to Kimbrough et al. (2008), fossil fuel and waste burning, mining and ore processing, 

chemical production, and agriculture are the sources largely responsible for the elevated metals 

concentrations observed in coastal waters. Metals are transported to coastal waters primarily from 

runoff and atmospheric deposition. The relative contribution from each source varies by metal, 

proximity to sources, and chemical phase (dissolved or particulate-bound).

Mercury, one of the most prevalent of the metals contaminants, is delivered to most aquatic 

ecosystems by deposition from the atmosphere, primarily through precipitation (USGS 1995). 

Anthropogenic sources of mercury contamination in the atmosphere include coal combustion, 

chlorine alkali processing, waste incineration, and metal processing (USGS 1995). Once deposited 

in aquatic ecosystems, mercury may be transformed to methylmercury; an organic form of 

mercury that is the most toxic and that readily bioaccumulates and is biomagnified (USGS 2000). 

Methylation occurs in sediments of freshwaters, estuaries and coastal zones; however the rate 

of net methylmercury creation from different sediment types is highly variable with higher 

methylation rates occurring in the organically rich, anoxic sediments common in wetland areas 

(Luoma and Rainbow 2008). Methylmercury has been identified as potent neurotoxicant (USEPA 

2001b). Factors such as chemical form, dose, route of ingestion, exposed organism’s species, 

sex, age, and overall health play a role in determining the level of toxicity and the environmental 

effects. Acute exposure to mercury most commonly affects the central nervous system and 

kidneys in fish, birds and mammals (USEPA 2011b).

The presence of ecological and biological effects resulting from metal contamination has been 

evident since the height of the Industrial Revolution; a substantial amount of literature related to 

metals in the environment dates as far back as the mid-1800s (Kapustka et al. 2004). Trace metal 

concentrations in the Chesapeake Bay peaked in the 1970s and have slowly declined since the 

1980s; despite this downward trend, concentrations remain elevated in some regions of the Bay 

(Hartwell and Hameedi 2007). 

Occurrence information for this chapter was derived primarily from the 2010 Water Quality 

Assessment Reports (Clean Water Act 303(d)/305(b) Integrated Reports) of each Bay jurisdiction. 

This chapter indentifies impairments and occurrences and does not attempt to compare loadings 

among jurisdictions. In addition to the assessment reports, information was captured from State 

fish consumption advisories, fish tissue monitoring program data from Maryland and Virginia, 

various reports from NOAA and other Federal agencies, as well as several studies related to metal 
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contamination throughout the Chesapeake Bay region (DDOE 2010, DNREC 2010, MDE 2010, 

PADEP 2010, VADEQ 2010, WVDEP 2010, Mason et al. 1999, TetraTech 2006, Hartwell and Hameedi 

2007, Fulton et al. 2007). 

Because in-depth review of the biological and ecological effects is beyond the scope of this 

report, a focus has been placed on selected metal contaminants that have occurred within the 

Chesapeake Bay Watershed. The exclusion of a particular metal in this discussion does not imply 

insignificance. Although the science surrounding metals and their biological and ecological 

effects continue to be studied intensively and is evolving rapidly, some aspects still lack sufficient 

information to support a quantitative assessment (USEPA 2007).

Water Column

Mercury

Mason et al. (1999) characterized mercury concentrations within the Bay (sites located in 

Baltimore Harbor, Potomac, Patuxent and Anacostia Rivers, and Maryland’s Bay waters). Overall 

mercury concentrations in the Chesapeake Bay were < 3 ng/L except for highly urbanized area 

sites such as the Baltimore Harbor and the confluence of the Anacostia and Potomac rivers. 

Total mercury concentrations in the upper Bay and in the subestuaries decreased down-estuary, 

coinciding with decreases in suspended particulate matter. Increased concentrations of mercury, 

especially methylmercury, were observed during seasonal bottom water anoxic conditions at the 

Baltimore Harbor. None of the Chesapeake Bay jurisdictions reported impairments for mercury in 

the water column.

Other Metals

Maryland, Pennsylvania and West Virginia each identified additional impairments for metals other 

than mercury. Both Maryland and West Virginia listed waters for aluminum and iron in the Bay 

watershed. West Virginia completed a TMDL addressing these impairments in the North Branch 

of the Potomac River Watershed (TetraTech 2006). Maryland identified additional water column 

impairments for copper and manganese. The copper impairment is isolated to Bodkin Creek in 

the Patapsco River watershed. The manganese, aluminum and iron impairments were located 

in the Upper North Branch of the Potomac River with acid mine drainage cited as a primary 

source (MDE 2010). Pennsylvania has completed more than 70 TMDLs for metals impairments, 

covering several hundred miles within the Bay watershed. The great majority of Pennsylvania’s 

metal impairments are for aluminum, iron, and manganese originating from abandoned mine 
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discharges in the coal areas of the upper Susquehanna watershed. Pennsylvania has a successful 

abandoned mine remediation program to address these water quality concerns.

Between 2008 and 2011 the Estuarine Probabilistic Monitoring Program (ProbMon) of the VADEQ 

collected clean dissolved trace metals samples from near surface waters at 182 estuarine sites 

in Virginia. Of these, 130 sites were within tidal portions of the Chesapeake Bay watershed. No 

exceedances of chronic saltwater standards were observed for any of the toxic metals evaluated 

(As, Cd, Cu, Pb, Hg, Ni, Se, Zn). Vanadium was added to the analyses beginning in 2010, and no 

exceedances of USEPA’s recommended saltwater criterion for vanadium (50 μg/L) were observed 

among the 58 samples evaluated from the Chesapeake watershed (D. Smith, VADEQ, personal 

communication).

Sediment

Mercury

None of the Bay jurisdictions documented sediment impairments for mercury. Concentrations 

of mercury in sediment in the Baltimore Harbor region have been found to exceed 1 ppb with 

methylmercury concentrations approaching 0.010 ppb (Mason et al. 1999). 

Other metals

Maryland identified sediment impairments specifically attributable to metals concentrations 

(chromium, lead and zinc). These metals impairments were localized to the Baltimore Harbor 

area, Bear Creek and Curtis Bay Creek. Zinc was the most widespread of the metals contaminants, 

affecting an area of approximately seven square estuarine miles. Sediment concentrations in 

excess of the State standards for the identified metal compounds have the potential to adversely 

affect aquatic life (MDE 2010).

Virginia conducted additional sediment monitoring as part of a toxicological characterization 

effort (Roberts et al. 2002, 2003, 2004). Sediment metals concentrations were below the ER-M 

with a few exceptions: in the Mattaponi, manganese exceeded the ER-M with a value of 3.38x106

ppb (Roberts et al. 2004). Various other metals were detected; however, no other metals 

exceeded the respective ER-M. In each study, the calculated SEM/AVS ratios (simultaneously 

extracted metals/acid volatile sulfides) indicated that the sediments may have additional metals 

binding capacity (Roberts et al. 2002, 2003, 2004). 
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Additional work on sediment in the Nanticoke identified detectable concentrations of all seven 

metals tested (arsenic, cadmium, chromium, copper, lead, mercury and zinc). Zinc concentrations 

were the highest, with values ranging from 91,100 – 224,000 ppb. None of the detected metals 

exceeded the respective ER-Ms (EA 2006).

In order to characterize the extent and magnitude of contaminated sediments in the Chesapeake 

Bay, Hartwell and Hameedi (2007) analyzed sediment samples for a suite of metals and metalloids: 

aluminum (Al), antimony (Sb), arsenic (As), cadmium (Cd), chromium (Cr), copper (Cu), iron (Fe), 

lead (Pb), mercury (Hg), nickel (Ni), silver (Ag), selenium (Se), tin (Sn), and zinc (Zn). The report 

documented metals concentrations above background at nearly all locations. The authors noted 

declining concentration trends with some site-specific exceptions, such as the Patuxent River 

and several southeastern tributaries that demonstrate high levels of cadmium. According to the 

report, Baltimore Harbor is one of the most contaminated sites in the Bay. Metals concentrations 

in sediment near Ft. McHenry, the Patapsco River, and the Potomac River’s Swan Point and Mattox 

Creek are in the top percentile for toxic metals nationwide. The Elizabeth River is also heavily 

contaminated; concentrations for most constituents are in the upper 75th percentile nationwide. 

In the upper part of the Bay, concentrations of metals in sediment are higher than in the middle 

and lower Bay, specifically for manganese, nickel, chromium and lead that are elevated by 

factors of 1.5 to 2 above background concentrations. Throughout the Bay copper and zinc are 

elevated by factors of 1.5 to 3.5 above background concentrations. The report cited the Fall Line 

Toxics Monitoring Report that calculated an enrichment rate of 110 times for cadmium in the 

Susquehanna River (CBP 1996). Sites near heavily developed areas, such as the Magothy River 

in Maryland and Broad Bay in Virginia Beach, also exhibited elevated concentrations of metals. 

Large western tributaries of the Bay had slightly higher concentrations than the lower mainstem. 

The authors attribute the high degree of variability in metals concentrations from site to site 

to interactions involving sediment grain size, proximity to sources, and the inherent particle 

reactivity of the elements. 

Fulton et al. (2007) inventoried toxic contaminants in five selected areas of the Bay for which 

data previously had been insufficient: the Chester River, Nanticoke River, Pocomoke River, lower 

Mobjack Bay (Poquosin and Back Rivers), and the South and Rhode Rivers. In the study, sediment, 

water column and benthic analyses were examined to characterize the biological effects of 

the contaminants. Total metal concentrations in sediment were found to be highest in the 

lower South River and the middle Chester River. Of the 60 stations, only two had contaminant 
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concentrations that exceeded the ERM for any analyte; in the upper South River concentrations 

at two stations exceeded the ERM (51.6 ppm) for nickel. The most common metal Effect Range 

Low (ERL) exceedances were arsenic, nickel and zinc in the Chester River, arsenic and nickel in the 

Nanticoke River, and arsenic in the Pocomoke River. 

Fish and Shellfish

Mercury

The primary method of human exposure to mercury is through ingestion of contaminated fish 

and shellfish (USGS 2000). States monitor levels in fin fish tissue and issue fish consumption 

advisories for the species that are found to have elevated levels in specific areas in order to 

protect human health. An advisory recommends a particular number and size of meals that 

can be safely consumed by the general public, pregnant women, and children; avoidance 

recommendations advise that no meals should be eaten regardless of meal size or population 

demographics. 

Maryland currently has fish consumption advisories due to mercury for over 10 species of fish in 

rivers, lakes and reservoirs within the Chesapeake Bay watershed. The number of recommended 

meals per month for the general population ranges from one to eight. The State of Maryland 

reported impairments for approximately 12,500 acres of impoundments. A TMDL report has been 

completed for nearly 8,200 of those acres (MDE 2010).

The Commonwealth of Virginia currently has fish consumption advisories for mercury for nine 

species in the Chesapeake Bay and small coastal basins, James River, Rappahannock River, and York 

River basins. The number of recommended meals for the general population is no more than two 

per month for all species. Virginia does not have any avoidance recommendations due to mercury 

in fish tissue at this time. The Commonwealth of Virginia reported a total of approximately 5,980 

acres in impoundments, approximately 300 miles in rivers and streams, and approximately 23 

square miles in estuaries as being impaired by mercury in fish tissue (VADEQ 2010). 

The District of Columbia advises the general population avoid carp, eel, and catfish from the 

Potomac and Anacostia Rivers; however the cause of advisory is not specified. No impairments 

specifically attributable to mercury in fish tissue have been reported in DC (DDOE 2010). 
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The State of New York notes impairment in the Lower Susquehanna River, as well as the Chenango 

River, a tributary to the Susquehanna. For these waters, the State advises that men over the age 

of 15 and women over the age of 50 consume no more than one meal per month of walleye 

longer than 22 inches. The State recommends that women and children eat none. A general fish 

advisory exists for all other fish of the Susquehanna which recommends that the public consume 

no more than four one-half pound meals a month of fish, with no more than one meal per week 

(NYDH 2012). 

The State of West Virginia has issued statewide advisories of one to two meals per month 

for more than 10 species within waters flowing to the Chesapeake Bay as a result of mercury 

contamination. There is a no consumption recommendation for carp in the Shenandoah 

River, and a limit of one meal per month for smallmouth bass longer than 12 inches. This 

recommendation is driven largely by PCB concentrations though mercury is an underlying 

contaminant. The State of West Virginia did not list any of the free flowing streams to the 

Chesapeake Bay as impaired for mercury in the 2010 Integrated Report (WVDEP 2010). 

The Commonwealth of Pennsylvania has issued advisories for 37 water bodies within the 

Susquehanna River basin due to mercury contamination; consumption recommendations are 

for one to two meals per month. One lake within the Potomac River basin has an advisory of two 

meals of walleye per month as a result of mercury contamination. In the 2010 Integrated Report, 

the Commonwealth of Pennsylvania reported that approximately 290 miles of rivers and streams 

within the Chesapeake Bay watershed are considered to be impaired due to mercury in fish tissue. 

Approximately 2,052 acres of impoundments are considered to be impaired for mercury in fish 

tissue (PADEP 2010). Pennsylvania is actively monitoring fish tissue mercury levels in the lower 

Susquehanna River (2006, 2008, 2010, 2011, 2012) and none of the results have resulted in a fish 

consumption advisory for mercury (P. Buckley, PADEP, personal communication).

The State of Delaware has a statewide advisory of no more than one meal per week for all waters 

of the state that are not covered by an explicit, waterbody-specific advisory. This advisory 

is meant to account for mercury and any other contaminants that might be present but not 

identified through a detailed study (R. Greene, DNREC, personal communication). 

The toxicity of mercury to fish is associated with methylmercury concentrations in fish tissue. 

Methylmercury is a highly toxic compound that crosses biological membranes, accumulates in 

the exposed species, and can be biomagnified up the food chain (Sandheinrich and Wiener 2011). 
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State monitoring programs frequently measure only total mercury, and criteria (such as USEPA 

2001b) are typically stated as criteria for methylmercury. Because there is little difference in the 

fish tissue concentrations when both compounds are measured, conversion from total mercury to 

methylmercury requires a decrease of 10% at most.

The USEPA (2001b) issued a water quality criterion for methylmercury of 0.3 ppm ww fish 

tissue for the protection of human health. This criterion was developed based on the basis of 

a review of high-dose exposures (i.e., Minimata Bay, Japan) that resulted in neurotoxic effects 

including mental retardation, cerebral palsy, deafness, blindness in utero and sensory and motor 

impairment in adults. Studies of chronic low level mercury exposure have shown developmental 

effects in children including abnormalities in memory, attention, and language. Other studies 

have reported cardiovascular and immunological effects. 

Within the Bay watershed, Virginia’s South River is the area of greatest concern with respect to 

mercury contamination. Except for stocked trout which spend little time in the river, no fish 

consumption is recommended. Between 1929 and 1950, mercury was used as a catalyst in 

fiber production at the DuPont plant in Waynesboro, Virginia. During that time, strict storage 

and disposal regulations did not exist, and mercury was released into the South River. A serious 

contamination problem was discovered in the 1970s (South River Science Team 2009); the VDH 

subsequently issued advisories for fish consumption along the South River and on the South Fork 

Shenandoah River. As part of a settlement agreement, DuPont paid a penalty and established 

a State-administered (by VADEQ) trust fund to support a 100-year mercury monitoring program 

for fish, water, and river sediment administered. The most recent fish sampling data listed on the 

VADEQ web site are from 2007 (VADEQ 2012c). At several sampling locations downstream from 

the plant, mercury concentrations frequently exceeded 1.5 ppm (five times the USEPA 2001b 

criterion) in smallmouth bass (Micropterus dolomieu), redbreast sunfish (Lepomis auritus), and 

white sucker (Catostomus commersoni). 

Sandheinrich and Wiener (2011) reviewed the toxic effects of methylmercury by evaluating 

the relation between tissue concentrations and toxic effects. Mercury exposure can adversely 

affect the survival, growth, and reproduction of fish. Wiener and Spry (1996) reported that 

mortality due to mercury poisoning is observed only at extremely high tissue concentrations 

(6 - 20 ppm wet weight in muscle). Behavioral effects include hyperactivity and altered shoaling 

activity (Webber and Haines 2003), which occurred in a laboratory study with golden shiners 
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(Notemigonus crysoleucas). The fish were fed a diet containing 0.96 ppm methylmercury and had 

whole body concentrations of 0.52 ppm wet weight. Reduced spawning success occurred in 

fathead minnows (Pimephales promelas) at whole body minus gonad concentrations of 0.71 to 

0.92 ppm wet weight (Drevnick and Sandheinrich 2003). Methylmercury causes oxidative stress 

in fish tissues through the formation of radical oxygen species and lipid peroxidation. Organs 

affected include the brain and liver (summarized in Sandheinrich and Wiener 2011). 

On the basis of their literature review, Sandheinrich and Wiener (2011) concluded that effects on 

fish biochemical processes, damage to cells and tissues, and reduced reproduction have been 

demonstrated at methylmercury concentrations of about 0.3 to 0.7 ppm in the whole body or 

0.5 to 1.2 ppm in muscle tissue. Given the high concentrations present in fish in Virginia’s South 

River, there is a great likelihood that fish in this water body are being adversely affected by 

methylmercury exposure. 

Other Metals and Metalloids

For the assessed waters, the states did not identify any impairments in fish tissue for metals other 

than mercury. According to the MDE human health screening study of finfish tissue collected 

from 1985 to 1997, tissue concentrations in finfish from the Maryland portion of the Chesapeake 

Bay did not exceed the established conservative human health risk-based screening values for 

arsenic, silver, cadmium, chromium, copper, mercury, nickel, and zinc (MDE, undated). Although 

concentrations of inorganic arsenic in a few individual bluefish, striped bass, and white perch 

exceeded the screening values, the MDE noted that those species exhibit migratory behavior 

and are therefore not considered to be a target species in the Core Monitoring Program (MDE, 

undated). It should be noted that this report is 15 years old; some screening values may be 

different now (2012). Additionally, the MDE report focuses on human health criteria only; aquatic 

life criteria are not examined. 

The Commonwealth of Virginia conducted a statewide fish tissue monitoring survey most 

recently in 2008. A review of the data for waters within the Chesapeake Bay watershed indicates 

that concentrations of arsenic in 83 of 477 samples from the Bay watershed exceeded the VADEQ 

screening value for arsenic (VADEQ 2012b). Although arsenic is detected at concentrations in 

excess of the screening value, VADEQ has not identified these waterbodies as impaired. This is 

in acknowledgement of the fact that the fish tissue data are based on the concentration of total 

arsenic, whereas toxicity is associated only with inorganic arsenic. Inorganic arsenic may be a very 
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small portion of the total arsenic value. VADEQ uses this information to identify areas in need of 

further study and has indicated that a method for measuring inorganic arsenic may be utilized 

for these efforts (VADEQ 2010). Concentrations in several samples exceeded the detection limits 

of <0.1 ppm ww for lead and <0.5 ppm ww for selenium (VADEQ 2008b). Because available data 

were insufficient to identify a threshold associated with a risk of human toxicity, these waters 

were not identified as impaired. Since 1986, NOAA’s Mussel Watch program has measured metal 

concentrations in oysters (Crassostrea virginica) within the Chesapeake Bay and its tributaries at 

five locations in Maryland and five locations in Virginia as part of a national effort (Kimbrough et 

al. 2008). Data collected in 2004-2005 from five locations in Maryland and five locations in Virginia 

were reported in an appendix of the report, with summary data for arsenic cadmium, copper, 

mercury, nickel, lead, tin, and zinc. For Maryland, data are reported on cadmium, copper, tin, and 

zinc which were rated as medium or high in at least one location in the national comparison. 

Cadmium concentrations ranged from 3.2 to 14 ppm, with all except the highest concentration 

described as medium in the national comparison. The two highest copper concentrations were 

301 ppm (rated as medium) and 867 ppm (rated as high). The two highest tin concentrations 

(0.25 and 0.44 ppm) were both rated as medium. Zinc concentrations of 2550, 2660, and 4570 

ppm were all rated as medium and 12,000 ppm was rated as high. In Virginia, concentrations 

of cadmium, copper, and zinc were rated as medium or high in the national comparison. The 

three highest cadmium concentrations were 4.8 and 4.8 (medium), and 10 ppm (high). The three 

highest copper concentrations were 285 and 308 (medium), and 1460 (high). The four highest 

zinc concentrations (2720, 3200, 3770, and 8110 ppm) were all rated as medium.

Wildlife

Mercury

Mercury concentrations in blood have been measured in several species of birds in the 

Chesapeake Bay watershed. Levels in osprey nestlings from regions of concern and reference 

sites ranged from 0.105 to 0.470 (Rattner et al 2008); adult tree swallow samples collected at 

a contaminated site in the South River (tributary of the Shenandoah River, which flows into 

the Potomac) had mean concentrations of <3.56 ppm ww (Brasso and Cristol 2008; Hawley et 

al. 2009), and values were < 0.5 ppm ww in Nelson’s sparrow (Ammodramus nelsoni) and the 

saltmarsh sparrow (A. caudacutus) from several Bay locations. With the exception of tree swallow 

samples from the South River, mercury concentrations observed in these species generally were 

less than the 0.400 ppm ww that has been associated with adverse sublethal effects in birds 

(Scheuhammer et al. 2007), although recent studies suggest that such concentrations could evoke 

endocrine disruptive effects in nestlings (Wada et al. 2009).
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Mercury is readily incorporated into feathers in growing birds and adults following molt. Mercury 

levels in feathers of 14- to 16-day-old nestling black-crowned night-herons from Baltimore Harbor 

and Holland Island did not differ, and ranged only from 0.04 to 0.23 ppm dw (Golden et al. 2003a). 

Mercury concentrations in feathers of 40- to 45-day-old osprey nestlings were greater in the 

Elizabeth River (0.26-2.40 ppm dw) compared to Baltimore Harbor and the Patapsco River, the 

Anacostia and the middle Potomac, and the South, West and Rhode Rivers (values ranged from 

0.01-1.53 ppm) (Rattner et al. 2008). Molted feather samples collected from 83 occupied bald 

eagle nests in the Chesapeake averaged 3.82 ppm dw (Cristol et al. 2012). Feathers from adult 

tree swallows in the vicinity of the South River in Virginia had average mercury concentrations of 

13.55 ppm dw (Brasso and Cristol 2008). With the exception of samples collected in the mercurycontaminated South River location in Virginia, values were generally well below the concentration 

commonly associated with adverse effects (7.5 ppm dw) in birds (Eisler 1987). 

Since the mid-1990s, several studies that examined eggs from ospreys (Rattner et al. 2004), 

peregrine falcons (USFWS et al. 2004; Clark et al. 2009), common terns (French et al. 2001) and 

bald eagles (Mojica and Watts 2008, 2011) indicated that mercury concentrations averaged well 

below 0.5 to 1.5 ppm ww, the generally accepted toxicity threshold for reproductive effects in bird 

eggs (Wiener et al. 2003). 

Other Metals and Metalloids

The Contaminants Exposure and Effects – Terrestrial Vertebrates (CEE-TV) database contains more 

than 50 records for Chesapeake waterbirds from 1988 to the present (2012) that describe lead 

concentrations in liver (Rattner and McGowan 2007; B. Rattner, USGS, unpublished data). For 26 

of these records (representing 84 individuals including waterfowl, geese, and bald eagles), lead 

concentration in liver exceeded 2 ppm wet weight, a value associated with subclinical poisoning 

in waterfowl (Franson and Pain 2011), and ranged up to 183 ppm dry weight in one of the bald 

eagles (> 10 ppm wet weight is compatible with death; Franson and Pain 2011). Many of these 

necropsy reports were related to the ingestion of spent lead shot that was historically used for 

hunting waterfowl but was phased out of use by 1991. Lead can be incorporated into feathers 

of growing birds (Golden et al. 2003b), and has been proposed as a sensitive minimally invasive 

indicator of exposure. Feathers of black-crowned night-heron nestlings from Chincoteague Bay 

and Holland Island were lower (<0.13 ppm dw) than feathers of 14- to 16-day-old nestlings from 

Baltimore Harbor (0.32 ppm) (Golden et al. 2003a). A similar tendency was noted in feathers 

of 40-45-day-old osprey nestlings collected from Baltimore Harbor in 2000 (geometric mean: 
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1.25 ppm dw) compared to nestlings from the South River reference area near Annapolis (0.66 

ppm), whereas those samples from the middle Potomac River watershed were intermediate (0.96 

ppm) (Rattner et al. 2008). Concentrations in osprey nestling feather samples collected from the 

Elizabeth River in 2001 (1.47 ppm dw) were significantly greater than the South, West, and Rhode 

Rivers reference site (0.54 ppm) (Rattner et al. 2008). Although lead concentrations in feathers of 

waterbirds in the Chesapeake Bay regions of concern are elevated compared to reference sites, it is 

not possible to relate the concentrations to toxicity thresholds or assess the risk of lead in feathers 

to overall waterbird health.

Reports of lead in liver of wild mammals in the Chesapeake Bay watershed are far more limited, and 

focused on military sites with known or suspected contamination. Lead was not detected in liver or 

kidney of woodchucks (Marmota monax) collected in the vicinity of small arms and skeet ranges at 

Aberdeen Proving Grounds in Maryland, although low levels were detected in blood (mean values 

< 0.92 ng/ml ww) and bone (mean values < 13 ppm dw) indicating its probable bioavailability 

(Johnson et al. 2004). 

The toxicity threshold for cadmium in liver and kidney has not been adequately established for 

birds (perhaps >100 ppm ww; Beyer 2000), and seemingly elevated concentrations in individuals 

may merely reflect that cadmium values increase with age. Cadmium is rarely detected in eggs, 

and poorly transferred into feathers of nestlings. Cadmium was not detected in blood of osprey 

nestlings collected in regions of concern, and concentrations in feathers were <0.28 ppm dw 

(Rattner et al. 2008). 

Selenium concentrations in common tern and peregrine falcon eggs averaged <2.6 μg/g dry 

weight (French et al. 2001; U.S. Fish and Wildlife Service et al. 2004; Clark et al. 2009; J.B. French, 

USGS-Patuxent Wildlife Research Center, unpublished data), and were clearly less than levels 

causing embryotoxicity, although this threshold is contentious (Ohlendorf 2003). Selenium was 

present in blood of nestling ospreys in the range of 2.34 to 18.2 ppm dry weight, and in feathers in 

the range of 0.71 to 9.73 ppm, neither of which approached levels associated with toxicity (Rattner 

et al. 2008).

Concentrations of aluminum, boron, chromium, copper, iron, manganese, strontium and zinc were 

often detected in tissues and eggs of Chesapeake Bay waterbirds collected since 1988, and were 

generally in the range of values for healthy captive birds. Arsenic, barium, beryllium, molybdenum, 

nickel and vanadium were commonly analyzed for but rarely detected in tissues and eggs (Rattner 

and McGowan 2007). 
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Conclusions

Adequate data to evaluate the extent and severity of metals and metalloids exists in the various 

media as a result of both Federal and routine jurisdiction monitoring. Mercury impairments 

dominate the 303(d) listings identified for metals with more than 600 river miles and 20,000 acres 

of impoundments listed for fish tissue impairments. Though impairments in the water column 

are not as common, both the States of Maryland and West Virginia have identified impairments 

for aluminum and iron. Pennsylvania has identified hundreds of miles of impairment for 

metals, nearly all attributable to abandoned mine discharges. In addition, the State of Maryland 

has identified sediment impairments for chromium, lead and zinc. The impact of mercury 

contamination on fish tissue is widespread based on documented fish consumption advisories. 

However, contamination from other metals is more localized, affecting limited areas of the Bay 

watershed. 
3.0 RESPONSES OF FISH TO CUMULATIVE AND INTERACTING STRESSORS

104

Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

In addition to cataloging the extent and severity of toxic effects in the Bay by contaminant class, 

there is a need to describe studies that indicate the adverse effects of fish exposure to complex 

mixtures of both legacy contaminants and contaminants of emerging concern (CEC). These 

complex mixtures can have additive (Silva et al. 2002; Brian et al. 2005; Correia et al. 2007; Vajda 

et al. 2008) as well as synergistic or antagonistic effects (Micael et al. 2007; Santos et al. 2006; 

Sárria et al. 2011). In addition, many chemicals, particularly those that affect the endocrine and 

immune systems, may not elicit the typical dose-response curves. Significant effects may be 

observed at very low levels, and nonmonotonic (nonlinear) dose responses are increasingly 

demonstrated (Welshons et al. 2003; Vandenberg et al. 2012). Additionally, most monitoring 

of chemical concentrations is a snapshot in time, and the results depend on sampling time and 

numerous environmental factors such as climatic conditions and flow. For instance, atrazine 

concentrations measured in the Monocacy River (Maryland) in the spring were 100 times those 

measured in the fall (Alvarez et al. 2009). Short-term exposures, particularly in early life stages, 

can have long-lasting effects on both the endocrine and immune systems (Leet et al. 2011; 

Milston et al. 2003; Mcallister and Kime 2003; Liney et al. 2005). 

For these reasons, biological effects monitoring is 

increasingly recognized as an important factor in 

assessing the cumulative and interacting effects 

of toxic chemicals as well as other stressors (such 

as low oxygen, temperature, and river flow) 

on ecosystem health (Dubé et al. 2006; Ankley 

et al. 2010). By collecting data at many levels 

of biological organization (sub-organismal to 

populations) along with systematic evaluation 

of water quality and other stressors, researchers 

can use these measurements as environmental 

indicators to document habitat quality and 

evaluate trends.

The following biological indicators of compromised fish health have been observed within the 

Chesapeake Bay watershed: increased incidence of infectious disease and parasite infestations 

contributing to increased mortality; feminization (intersex, plasma vitellogenin) of largemouth 

and smallmouth bass and other signs of endocrine disruption; reduced reproductive success 

Mouth lesion on a brown bullhead from the South River, Maryland, later 

diagnosed as squamous cell carcinomas. Photo by Fred Pinkney, USFWS
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and recruitment of yellow perch in certain highly urbanized tributaries; and tumors in bottomdwelling fish. In this section of the report, we describe these effects, and the evidence that 

suggests possible associations with exposure to toxic chemicals, so that they are considered in 

developing goal setting recommendations for the CBP. 

Infectious Diseases, Parasites, Immune Suppression, and Fish Kills 

Worldwide, there have been substantial increases in disease reports for freshwater fishes, 

amphibians and crayfish from 1970 to 2009 (Johnson and Paull 2010) and anthropogenic 

drivers, including toxic chemicals may play direct and indirect roles in these increases. Recent 

studies suggest increased disease incidence can be related to human-induced land use changes 

(Patz et al. 2004), increased nutrient concentrations (McKenzie and Townsend 2007), climate 

change (Karvonen et al. 2010), and toxic chemicals (Feingold et al. 2010). Exposure to a range 

of chemicals has been shown to cause immune system effects that can increase susceptibility, 

increase the persistence and hence shedding of infectious organisms, and alter the severity of 

disease (e.g., Arkoosh et al. 1998; Ross 2002; Zelikoff et al. 2002). Hence, understanding disease 

epizootics in wild populations is complex and requires a multidisciplinary approach. 

A number of epizootic diseases have been documented in the Chesapeake Bay in recent 

decades. Ulcerative skin lesions of menhaden (Brevoortia tyrannus) were determined to be 

caused by the oomycete, Aphanomyces invadans (Blazer et al. 1999, 2002; Kiryu et al. 2002), 

a pathogen responsible for major mortalities in freshwater and estuarine fishes worldwide 

(Baldock et al. 2005). Skin and internal lesions have also been observed in a high percentage (> 

50%) of striped bass (Morone saxatilis) from the Bay and its tributaries. Disease in striped bass 

has been attributed to a variety of Mycobacteria spp. (Ottinger and Jacobs 2006; Gauthier and 

Rhodes 2009). Mycobacteriosis is a chronic disease; however, negative population-level effects 

have been demonstrated indicating the importance of disease in management and stock 

assessment (Gauthier et al. 2008). The role of toxic chemicals in susceptibility to these infectious 

diseases is currently unknown. 

Within the Potomac River watershed, in 2002, major mortality of smallmouth bass (Micropterus 

dolomieu) and other freshwater fishes was observed in the South Branch Potomac River. Similar 

fish kills occurred in the North Fork of the Shenandoah River in 2004, the South Fork Shenandoah 

River in 2005 and the Monocacy River in 2009. Adult fishes, during the spring were affected 

and most dead and dying fish exhibited a variety of external skin lesions. A number of potential 
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pathogens, including the bacteria Aeromonas salmonicida, Flavobacterium columnare, Aeromonas 

hydrophila and other motile Aeromonads, internal trematode and myxozoan parasites, as well 

as external parasites including trematodes and leeches have all been identified or cultured. 

However, no consistent pathogen or parasite was identified as a single cause of the mortalities 

at all sites (Blazer et al. 2010). These findings suggest certain sensitive (immunocompromised) 

species are affected by a variety of opportunistic pathogens eventually leading to lesions and/or 

death. 

Raised skin lesions which microscopically range from hyperplasia (proliferation of normal 

epidermal cells) to papillomas (benign neoplasia) were observed in adult bass from the Potomac 

(Blazer et al. 2010) and Susquehanna Rivers (V. Blazer, USGS, personal communication). The cause 

of these is currently unknown. However, similar lesions have been described in a variety of fish 

species. In some instances, viral particles have been observed in similar lesions by using electron 

microscopy (Anders and Möller 1985; Quackenbush et al. 2001). In other cases, investigators 

have tried to culture and/or visualize viruses using electron microscopy and have not found any 

indication of a viral etiology. There are a number of reports of increased prevalence of these 

types of papillomas at sites impacted by industrial and/or sewage effluent (Korkea-aho et al. 

2006, 2008) and also by exposure to androgens (Kortnet et al. 2003). The International Council 

for the Exploration of the Sea (ICES) uses papilloma prevalence as one of suite of indicators to 

monitor environmental conditions (Bucke et al. 1996).

All of the Potomac River basin watersheds affected with fish kills were also areas where a 

high prevalence and severity of intersex or testicular oocytes was documented (see below; 

Blazer et al. 2007a, 2012). The co-occurrence of skin lesions/mortalities and signs of endocrine 

disruption further suggests that exposure to chemical contaminants and other stressors may 

contribute to the reduced health of these populations. For instance, whereas exposure to 

estrogenic compounds, is considered a primary cause of feminization (testicular oocytes and 

plasma vitellogenin) of male fishes, estrogens have also been shown to modulate disease 

resistance (Iwanowicz and Ottinger 2009; Robertson et al. 2009). Two other chemicals, atrazine 

and arsenic, known to affect disease resistance were identified in the fish kill and endocrine 

disruption research in the Potomac River basin (Blazer et al. 2010; Alvarez et al. 2008, 2009). 

Exposure to sublethal concentrations of atrazine, a commonly detected herbicide in the 

Chesapeake watershed, including sites where fish kills occurred (Alvarez et al. 2008, 2009), has 

been shown to increase susceptibility of silver catfish to Aeromonas hydrophila infection (Kreutz 
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3.0 RESPONSES OF FISH TO CUMULATIVE AND INCREASING AND INTERACTING STRESSORS

et al. 2010). Atrazine exposure is also associated with increased trematode infections in some 

amphibians (Rohr et al. 2008a, 2008b). Arsenic, used in pesticides as an additive in poultry feed 

and found naturally, has been reported to modulate the immune response of fishes (Hermann 

and Kim 2005; Lage et al. 2006) and is also associated with skin lesions in humans (Kazi et al. 

2009). Exposure to arsenic was shown to enhance the ability of A. hydrophila to colonize and 

disseminate within exposed catfish (Datta et al. 2009) and inhibit the ability of zebrafish (Danio 

rerio) to clear viral or bacterial infections (Nayak et al. 2007). A significant increase in arsenic 

concentrations in the skin and anterior kidney (major immune organ of fish) between March and 

May (leading up to the fish kills) was demonstrated in smallmouth bass from the Shenandoah 

River (Blazer et al. 2010). 

Feminization of Male Bass

Two indicators of feminization of male fishes that have been used worldwide are testicular 

oocytes (intersex) and vitellogenin. Experimental exposures of various fish species to natural and 

synthetic estrogens have shown that the most sensitive stage for induction of oocytes within 

the testes is during sexual differentiation, within two to three weeks after hatching. Exposures at 

these early life stages can lead to a greater sensitivity to estrogenic exposures later in life (Liney 

et al. 2005). Plasma and liver vitellogenin mRNA may be indicative of more recent exposure. 

Estrogens induce vitellogenin production by the liver. Normally this process occurs only in eggproducing females; however, exposure to abnormal concentrations of estrogens can induce 

immature or male fishes to produce vitellogenin.

Studies in the Chesapeake Bay watershed have concentrated on smallmouth and largemouth 

(M. salmoides) black bass species, although species comparisons have been conducted in some 

areas. Results of the U.S. Geological Survey Biomonitoring of Environmental Status and Trends 

program showed black bass to be sensitive species for reproductive endocrine disruption 

studies. In surveys conducted from 1995 to 2004 in nine river basins (Mississippi, Columbia, 

Rio Grande, Colorado, Yukon, Pee Dee, Apalachicola, Savannah, and Mobile), intersex was 

most prevalent in smallmouth and largemouth bass (Hinck et al. 2009). Largemouth bass 

were collected at 52 of the 111 sites sampled and at least one male with testicular oocytes was 

collected at 23 (44%) of these sites, with an overall prevalence of 18%. Smallmouth bass were 

collected at 16 sites and at 7 (44%) at least one male with testicular oocytes was observed, with 

an overall prevalence for smallmouth bass of 33% (Hinck et al. 2009). Similarly, largemouth bass 

within the Potomac River drainage also show signs of feminization, but may be less sensitive 
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3.0 RESPONSES OF FISH TO CUMULATIVE AND INCREASING AND INTERACTING STRESSORS

than smallmouth bass (Blazer et al. 2007; Iwanowicz et al. 2009). The prevalence of intersex in 

smallmouth bass in the Potomac River is considerably higher than that observed in the national 

survey, ranging from 50 to 100% (Blazer et al. 2007, 2012). 

With respect to reproductive endocrine disruption, most studies to date have focused on the 

prevalence and severity of, and factors that contribute to feminization (intersex and plasma 

vitellogenin in males). Because aquatic organisms may be exposed to complex mixtures of 

chemicals that can have additive, synergistic, or antagonistic effects, it is difficult to identify 

one chemical or one source of the estrogenic compounds. Hence, biological effects, as well as 

chemical monitoring, were used to determine the extent of reproductive endocrine disruption 

throughout the Potomac River and nearby drainages, to identify potential causes, sources, and 

associated land-use practices and to document effects. Initial biological effects monitoring (Blazer 

et al. 2007) indicated that the prevalence of testicular oocytes varied from 100% at some sites in 

the Shenandoah River (Virginia) to low or background levels (10-14%) at selected out-of-basin 

sites such as the Gauley, Tygart, and Cheat Rivers (West Viginia). Prevalence in the South Branch of 

the Potomac River in West Virginia was intermediate, with 50% to 75% of the male bass affected. 

These surveys indicated a probable gradient of intersex in West Virginia and Virginia associated 

with human population and agricultural land use in the counties containing the sites. Although 

sample sizes of male smallmouth bass were small at some sites, there was an indication that, even 

within out-of-basin rivers such as the Greenbrier River, prevalence of intersex at downstream sites 

increased with human population and agricultural land use (Blazer et al. 2007). 

Feminization of male fishes, in other parts of the world, has most commonly been associated with 

exposure to human wastewater-treatment plant (WWTP) effluents or other point sources. Hence, 

an early study in the Potomac River watershed to identify chemical compounds and associated 

point sources focused on sites upstream and downstream from WWTPs on the Monocacy River 

and Conococheague Creek, Maryland (Alvarez et al. 2009; Iwanowicz et al. 2009). Although 

some biomarkers such as gonadosomatic index and female plasma vitellogenin concentrations 

were adversely affected downstream from these WWTPs, the prevalence of males with testicular 

oocytes or plasma vitellogenin was not consistently higher downstream. A similar lack of increased 

prevalence of intersex downstream from WWTPs was found in the Susquehanna River drainage in 

Pennsylvania (V. Blazer, USGS, personal communication). Whereas prevalence was not significantly 

different, severity increased downstream, indicating that other sources may be inducing 

intersex early in development but that WWTP effluent may contribute to increased severity. The 
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3.0 RESPONSES OF FISH TO CUMULATIVE AND INCREASING AND INTERACTING STRESSORS

Pennsylvania project also demonstrated that white sucker (Catostomus commersonii) collected at 

the same sites as smallmouth bass had no signs of testicular oocytes, but male suckers did show 

vitellogenin. 

These findings led to investigation of other sources such as agriculture. A spring spawning study 

(Blazer et al. 2012; Kolpin et al. 2013) was conducted in which fish were collected just prior to 

spawning for a suite of reproductive endpoints. Sediment samples were collected from the 

smallmouth bass nests, integrative samplers were placed near nesting areas to bracket spawning, 

and discrete water samples were collected at the time of fish sampling. Because many of the 

chemicals currently of concern have biological effects at very low levels–commonly close to or 

below analytical capabilities–two approaches were used to assess water samples in conjunction 

with the biological effects in the fish. First, integrative passive samplers were deployed for 

approximately 30 days at the various sites, allowing the capture not only of cumulative base flow 

but also runoff events that may occur during the same time period. Second, extracts of discrete 

water samples as well as the integrated samples were tested for total estrogenicity. Results of 

the biological effects monitoring over two spring spawning seasons demonstrated a reduction 

in sperm count and in the motility of the sperm in bass from the Potomac River with respect to 

those at the reference site in the Gauley River, and an inverse relation between testicular oocyte 

severity and sperm motility. Second, a seasonal difference in intersex and vitellogenin in male bass 

was found. Third, for the various land-use characteristics (human population density, number of 

WWTPs, WWTP flow, and various agricultural attributes), intersex prevalence was associated only 

with percent agriculture and animal density in the catchment above the site. Intersex severity was 

associated with these two variables, as well as total animal feeding operations (AFOs), number of 

poultry houses, and WWTP flow (Blazer et al. 2012), also indicating that factors associated with 

agricultural runoff may be responsible for induction of intersex in these areas, whereas multiple 

sources increase intersex severity over the life of the fish.

Chemical analyses results obtained during the spawning study indicated that 39 of a total of 201 

target chemicals were detected in at least one discrete water sample. Atrazine, its degradate, 

and simazine (herbicides), iso--chlorotetracycline (antibiotic) and caffeine (stimulant) were the 

most frequently detected. In contrast, 100 compounds, including four biogenic hormones were 

detected in extracts from the integrated samplers. The concentration of atrazine in the discrete 

water samples and total hormone and sterols in the sediment were significantly related (positively 

correlated) with the prevalence and severity of testicular oocytes (Kolpin et al. 2013).
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Ciparis et al. (2012) conducted a landscape-scale study in smaller watersheds of the Shenandoah 

River basin, assessing effects of WWTP and AFOs on nutrients and estrogenic activity, using the 

bioluminescent yeast estrogen screen. Eighteen sites representing a gradient of influence from 

AFOs combined with presence/absence of WWTP discharge were sampled during both low and 

high flow periods. Strong relationships were found between watershed densities of AFOs and 

estrogen equivalents at all sampling times. The authors concluded that there was no consistent 

relationship between estrogenicity and WWTP discharges. 

Reproductive Impairment of Yellow Perch

Several Chesapeake Bay tributaries, including the Severn and South Rivers, were closed to 

yellow perch fishing for nearly 20 years. Despite both recreational and commercial fishing 

regulations, a continued lack of recruitment has occurred. Hence, the streams were reopened 

for recreational fishing in 2009. The rationale for reopening the Severn, South, and West Rivers 

was that these watersheds are extremely degraded habitat, reproductive output is low, and, 

therefore, recreational harvest would not affect the total reproductive capacity (Maryland 

Fisheries Service 2012). 

Larval presence, Lp, is defined as the proportion of 0.5-meter plankton tows with larvae during 

the peak weeks from late March through early May, and is used as an indicator of year class 

strength. The Lp index integrates egg production, egg hatching success, and survival of firstfeeding larvae. Brackish systems with small watersheds and high levels of development (South, 

Severn, and Magothy Rivers) have exhibited a persistent depression in Lp to below a reference 

minimum since 2002. Regression analyses indicated that development (percent impervious 

surface in the watershed--i.e., pavement, rooftops, and compacted soils) was negatively related 

to L

p

. Other systems may exhibit wide variation in larval presence, but low levels similar to 

those seen in these urbanized subestuaries are not common; however, since 2008 spawning 

site loss for anadromous fishes is evident in the urbanizing Mattawoman Creek watershed 

(Uphoff et al. 2011). 

Despite numerous studies, the contributing factors to the lack of reproductive success have not 

been identified. Neither excessive adult mortality nor reduced growth has been observed in 

Severn River yellow perch. Severn River broodstock induced to spawn in the hatchery produced 

visually abnormal egg chains and too few viable eggs to support hatchery production. Hence, 

the depressed egg and larval survival are hypothesized to be critical factors suppressing resident 
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3.0 RESPONSES OF FISH TO CUMULATIVE AND INCREASING AND INTERACTING STRESSORS

yellow perch populations in western shore subestuaries whose watersheds are subject to high 

levels of development (Uphoff et al. 2005). In recent surveys that assessed a suite of bioindicators, 

a number of biological effects were found in yellow perch from the Severn and South Rivers and, 

to a lesser degree, Mattawoman Creek. An apparent lack of final maturation, abnormal yolk, and 

zona pellucida (egg coat) were noted in ovaries collected during the spawning run. In the males, 

abnormal proliferation of Leydig cells was noted (Blazer et al. in press).

Skin and Liver Tumors in Bottom-Dwelling Fish

Although liver tumors have been clearly associated with sediment exposure to polycyclic 

aromatic hydrocarbons (PAHs) in bottom-dwelling fish (see Section 2.3), it is likely that exposure 

to other chemicals contributes either as an initiator or a promoter. Evidence that polychlorinated 

biphenyls (PCBs) and estrogens are believed to promote liver tumors is summarized in Sections 

2.1 and 2.9, respectively. 

In contrast, the causes of skin tumors in brown bullheads are more uncertain and the subject of 

intense investigation. These tumors are commonly grotesque in appearance and alarming to the 

public. From 1992 to 2008, the USFWS conducted surveys in the following low salinity waters of 

the Chesapeake Bay watershed: Potomac River (Quantico Embayment), Anacostia River, Rhode 

River, South River, Severn River, Back River, Furnace Creek, Farm Creek, Marumsco Creek, Neabsco 

Creek, Choptank River, Little Blackwater River, and Tuckahoe River (Pinkney et al. 2009, 2011). An 

early goal of the tumor surveys was to systematically evaluate tumor prevalence in bullheads 

from the Anacostia River in Washington, DC, where there was anecdotal evidence of fish with 

visible mouth lesions. The linkage between liver tumors and PAHs in this river is described 

in Section 2.3, as are studies relating liver tumors and PAH exposure in mummichogs in the 

Elizabeth River. 

A second goal of the bullhead surveys has been to understand the relation between skin 

tumors and exposure to environmental contaminants. High prevalence of skin tumors has been 

observed in the Anacostia, South, Severn, Little Blackwater, and Neabsco Rivers (Pinkney et al. 

2009, 2011). The skin tumor prevalence in the South River, in particular, attracted widespread 

attention as a result of the grotesque appearance of several individuals. Results of the first 

survey in 2005, revealed a 53% prevalence of skin tumors (Pinkney and Harshbarger 2005), as 

high as any reported in North America. A follow-up study was conducted with an archived 

sample from 2004 and multiple samples from 2007 and 2008, as well as samples collected from 
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3.0 RESPONSES OF FISH TO CUMULATIVE AND INCREASING AND INTERACTING STRESSORS

the nearby Rhode and Severn Rivers (Pinkney et al. 2011). Skin tumor prevalence in the South 

River bullheads ranged from 19 to 58%. Prevalence of liver tumors was low in most collections, 

ranging from 0 to 6% in all collections but one, in which prevalence was 20%. As part of the 

study, biliary PAH-like metabolites as well as polycyclic aromatic compound (PAC)-DNA adducts 

in skin and liver tissues were measured. The authors also investigated the concentrations of 

alkyl-DNA adducts, which arise from interactions of alkylating agents such as nitrosamines 

with DNA. South River sediment total PAH concentrations averaged only 3.6 ppm within two 

kilometers of the collection site; concentrations from the other rivers were smaller. Neither 

the sediment PAH concentrations, the PAH-like metabolite data, nor the PAC-DNA adduct data 

show a linkage between exposure to PAHs and liver or skin tumors in these rivers. The alkylDNA adduct data also were not consistent with liver- or skin-tumor prevalence (Pinkney et al. 

2011). Therefore, no chemical class has been associated with the high skin-tumor prevalence in 

the South River. On a broader scale, Pinkney et al. (2009) produced scatterplots of skin tumor 

percentages in relation to sediment total PAH concentrations near the Chesapeake Bay bullhead 

collection sites and found no relationship between prevalence and concentrations. 

A third goal was to use logistic regression to statistically analyze the Chesapeake Bay bullhead 

tumor database to determine which covariates must be considered when comparing tumor 

prevalence among collections and locations. Based on 1992 through 2008 Chesapeake bullhead 

tumor data, Pinkney et al. (2009) reported that length and sex (with females having a higher 

ratio) were significant covariates for liver tumors, and length was a significant covariate for skin 

tumors. This statistical analysis, enhances the utility of bullhead tumors as an environmental 

indicator as defined by USEPA (2006), since it allows researchers to make unbiased comparisons 

over space and time. 

Both the bullhead and mummichog tumor surveys meet the criteria proposed by USEPA 

(2006) to select environmental indicators. The use of mummichog liver histopathology as an 

environmental indicator, especially in the Elizabeth River Program monitoring studies (Chapter 

2.3, e.g., Vogelbein and Unger 2011), make it a useful complement to the bullhead for higher 

salinity waters of the Chesapeake Bay watershed. 
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4.0 CONTAMINANT EXPOSURE AND RESPONSES IN WILDLIFE

Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

This section compiles and summarizes wildlife information in Chapter 2 to provide an overall 

assessment of contaminant exposure and responses in wildlife in the Chesapeake Bay 

watershed. The vast majority of the studies focus on birds.

On the basis of earlier reviews (Ohlendorf 

and Fleming 1988; Heinz and Wiemeyer 1991; 

Rattner and McGowan 2007) and the present 

analysis, tissue concentrations of organochlorine 

pesticides and their metabolites have declined 

and widespread adverse reproductive effects 

on Chesapeake Bay waterbirds have subsided. 

Nevertheless, in a few locations concentrations 

of organochlorine pesticides remain elevated, 

eggshell thinning associated with p,p’-DDE is 

apparent, reproduction may be impaired, and 

organochlorine pesticide-linked lethality has even 

been observed in high-trophic-level predatory 

birds (reviewed in Rattner and McGowan 2007). 

In numerous instances, organophosphorus and 

carbamate pesticides have been linked to avian 

die-off events in agricultural regions of the 

Chesapeake Bay watershed (Rattner and McGowan 

2007), although use of the most hazardous 

anticholinesterases (e.g., granular formulations of carbofuran and diazinon, parathion) has 

been curtailed. Residue data in wildlife are not available for many categories of pesticides 

(e.g., herbicides, fungicides, rodenticides), likely because most have very low potential to 

bioaccumulate and have short environmental half-lives. The toxicological effects of chronic lowlevel exposure to newer pesticides and mixtures are unknown.

Concentrations of PCBs in tissues of many species of Chesapeake Bay wildlife have not declined 

since the final USEPA rule restricting the manufacture, processing, and distribution became 

effective in 1979 (Rice et al. 2003). In some urbanized regions, exposure to PCBs appears to be 

substantial and molecular effects are apparent; these compounds may contribute to localized 

reproductive problems (Rattner et al. 1997; Rattner and McGowan 2007; Mojica and Watts 
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4.0 CONTAMINANT EXPOSURE AND RESPONSES IN WILDLIFE

2008, 2011). Concentrations of dioxins and dibenzofurans, which evoke toxicity by a similar 

mechanism to coplanar PCB congeners, are low in peregrine falcon eggs (Clark et al. 2009), but 

limited sampling-location data and the absence of exposure information in other species do not 

permit Bay-wide conclusions. 

Moderate to high concentrations of polybrominated diphenyl ether flame retardants (PBDEs) 

have been detected in eggs from predatory birds in the Chesapeake Bay watershed, and 

approach the lowest-observed-adverse-effect level for pipping and hatching success (Rattner 

et al. 2004; Chen et al. 2010). Because of their documented persistence and tendency to 

bioaccumulate, along with concerns about potential toxicity, certain penta- and deca-BDE 

formulations were phased out in 2004 and deca-BDE formulations will be phased out by 

the end of 2013 (USEPA 2012a). Exposure and effects data for other groups of new flame 

retardants (e.g., hexabromocyclododecane and organophosphate flame retardants including 

tris(3,5-dimethylphenyl)phosphate and tris(1,3-dichloro-2-propyl)phosphate) are lacking for 

Chesapeake Bay wildlife. Some exposure monitoring has been undertaken with surfactants, 

such as alkylphenols and ethoxylates, and concentrations of perfluorinated compounds were 

found to be low, and seemingly well below effect thresholds in wildlife (Rattner et al. 2004). 

On average there are about 500 reportable oil spill events in the Chesapeake Bay annually, but 

no major spill events have resulted in large-scale wildlife die-offs. Small spills and seepage 

of oil from marine vessels, and possibly pump facilities and industrial sites, are thought to 

be pervasive. The effects of chronic exposure to aliphatic and aromatic hydrocarbons in 

Chesapeake Bay wildlife have not been monitored, and are unknown.

Despite nationwide interest and concern about potential effects of household products, 

personal care products, and pharmaceuticals released from wastewater-treatment plants, 

septic systems, combined sewer outflows, and landfills, there have been no studies examining 

exposure and potential effects of these chemicals on Chesapeake Bay wildlife. Efforts are 

currently underway that are examining a suite of compounds (analgesics, antibiotics, antiinflammatories, antihypertensives, antihistamines, antilipemics, antiseizures, anticoagulants, 

stimulants, psychostimulants, parasiticides, antidepressants, and antiseptics) in blood of osprey 

nestlings (Lazarus et al. 2012). The extent of exposure and potential effects of these compounds 

on wild birds is as yet unknown.
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4.0 CONTAMINANT EXPOSURE AND RESPONSES IN WILDLIFE

Although use of lead shot for hunting waterfowl was phased out by 1991, lead exposure and 

its resulting adverse effects are not uncommon in the Chesapeake Bay watershed, particularly 

in industrialized regions. However, exposure to other metals, including cadmium, mercury, and 

arsenic, do not seem to be a significant stressor to wildlife in most of the Chesapeake Bay and its 

tributaries (Rattner and McGowan 2007).

Interest in the effects of biotoxins, particularly those associated with harmful algal blooms 

(HABs), on wildlife has increased. Since 2001, a protracted series of heron die-offs that may be 

linked to microcystins, a potent hepatotoxin from cyanobacteria found in harmful algal blooms, 

has occurred (Rattner et al. 2006). Global climate change could cause an increase in HABs in the 

Chesapeake Bay watershed, and adverse effects on waterbird health may be a prominent issue 

in the foreseeable future (O’Neill et al. 2012). 
5.0 SUMMARY AND CONCLUSIONS

116

Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

Summary of Extent and Severity of 

Occurrence of Contaminants 

Based on the information in Chapter 2, ten 

contaminant groups are summarized in terms 

of extent and severity.

Extent: Extent was characterized as 

“widespread, localized, or uncertain” and 

depended on a series of considerations using 

readily available information. If information 

on the occurrence of a contaminant group 

was acquired, an assessment of local versus 

widespread extent was made. However, if 

information was limited, the assessment of 

extent could only be classified as localized or 

uncertain.

» Widespread extent: For polychlorinated biphenyls (PCBs), polycyclic aromatic hydrocarbons 

(PAHs), herbicides (primarily atrazine, simazine, metolachlor, and their degradation products), 

and mercury, available information indicates widespread extent of occurrence throughout the 

Bay watershed. 

» Localized extent: For dioxins/furans, petroleum hydrocarbons, some chlorinated insecticides 

(aldrin, chlordane, dieldrin, DDT/DDE, heptachlor epoxide, mirex), and some metals 

(aluminum, chromium, iron, lead, manganese, zinc), available information indicates localized 

extent of occurrence. 

» Uncertain extent: For pharmaceuticals, household and personal-care products, 

polybrominated diphenyl ether (PBDE) flame retardants, some pesticides, and biogenic 

hormones, available information is insufficient to determine extent of contamination. 

However, the widespread distribution of known sources of these contaminants (e.g., 

wastewater effluents, agricultural runoff, etc.) in the watershed and the summarized 

occurrence data indicate that some contaminants from each of these groups may have the 

potential to be found in many locations throughout the Bay watershed. 
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5.0 SUMMARY AND CONCLUSIONS

Severity: Contaminant groups (or constituents) having impairments identified at many locations 

were considered widespread. Contaminants with impairments in a few locations are classified as 

having localized severity. All other contaminant groups are classified as uncertain.

» Widespread severity: For PCBs and mercury, impairments have been identified in many 

locations in the watershed, largely in response to concentrations in sediments and in fish 

tissues that frequently result in the need for fish-consumption advisories. 

» Localized severity: For dioxins/furans, PAHs, petroleum hydrocarbons, some chlorinated 

pesticides (aldrin, chlordane, dieldrin, DDT/DDE, heptachlor epoxide, mirex), and some 

metals (aluminum, chromium, iron, lead, manganese, zinc), the report identifies localized 

severity on the basis of impairments in a limited number of areas in the Bay watershed. 

» Uncertain severity: For atrazine, some pharmaceuticals, some household and personal-care 

products, some PBDEs, and biogenic hormones, severity as defined in this report could not 

be assessed. However, recent peer-reviewed research has documented sublethal effects for 

some compounds at environmentally relevant concentrations, raising concerns about the 

potential for adverse ecological effects. 

The availability of Chesapeake watershed monitoring data and established toxicity benchmarks 

is variable for the toxic contaminant groups assessed in this report. For some pollutants 

such as PCBs, and mercury, abundant data indicate extensive environmental distribution 

at concentrations that result in impairments and fish consumption advisories. PAHs and 

herbicides are also widespread throughout the system. Impairments for PAHs exist in some 

jurisdictions and therefore the severity was defined as localized. For herbicides severity, as 

defined in this report, is classified as uncertain although the published literature has reported 

effects on organisms exposed to concentrations detected in the watershed. For an additional 

group of pollutants-dioxin and furans, petroleum hydrocarbons, chlorinated insecticides and 

some metals-localized areas of contamination have been identified at concentrations resulting 

in impairment designations. For other contaminant groups such as certain pharmaceuticals, 

household products, personal care products, flame retardants and biogenic hormones, there 

are data available that suggest that these contaminants are in the Bay system; however, more 

monitoring data and effects research would be needed to assess the severity of effects. Finally, 

there are groups of toxic contaminants such as certain pesticides and pharmaceuticals for which 

monitoring data gaps exist. 
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5.0 SUMMARY AND CONCLUSIONS

The assessment made through this report has noteworthy limitations. It relies on available 

data from state and Federal environmental monitoring, most of which is short term in nature 

and does not allow for temporal assessment of trends. While the report focused mainly on 

data specific to Chesapeake Bay published after 2000, peer-reviewed literature that included 

data from outside the Chesapeake Bay watershed was used to provide additional perspective 

for contaminants that were either not monitored in the Chesapeake Bay watershed or 

where monitoring data were spatially insufficient. The cost of analytical tests for detection, 

quantification, and biological effects of toxic contaminants is significant and, therefore, the 

spatial coverage of the data is limited even where monitoring programs exist. The method that 

underlies the report’s conclusions is single-stressor oriented (or groups of chemically similar 

stressors such as congeners of PCBs) and does not consider the risk presented by additive, 

synergistic, or antagonistic effects of multiple chemical, biological, or physical stressors, nor 

does it address the growing body of toxicological research on sublethal effects. The latter 

is a major limitation to present paradigms in adequately addressing the severity of toxic 

contaminants in the environment. Where appropriate, references to this limitation have been 

discussed and relevant published research has been cited.

In conclusion, sufficient assessment and occurrence data were available to characterize a limited 

number of contaminant groups assessed as “widespread” or “local” in extent and severity. For 

these contaminant groups, this report can be used by the CBP partnership and others as an 

information resource to help inform decisions regarding whether to establish reduction goals 

and strategies. For other contaminants, further monitoring and research would better inform 

the understandings of the extent and(or) severity of effects. The summary table below (Table 2) 

illustrates these conclusions.
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5.0 SUMMARY AND CONCLUSIONS

Table 2. Summary of extent, severity, and monitoring and research gaps for specific 

contaminant groups in the Chesapeake Bay watershed.

Quantity of Information 

(available throughout 

the watershed or 

limited)

Geographic Extent 

(localized, widespread, or 

uncertain)

Severity (localized, 

widespread, or uncertain)

Monitoring and 

Research Gaps

Polychlorinated Biphenyls

Information available 

throughout the 

Chesapeake Bay 

watershed

Widespread. Detected in 

fish tissue and sediment. in all jurisdictions and 

Widespread. Impairments 

multiple water bodies 

in most jurisdictions are 

subject to fish consumption 

advisories. 

None* (see footnote)

Dioxins and Furans

Limited information to 

define extent. 

Localized. Detected in 

Elizabeth River, North 

Branch of the Potomac 

(WVA and MD) and in C&D 

Canal (DE). 

VA (Elizabeth River) and DE 

(C&D Canal) based on fish 

consumption advisories.

Localized. Impairments in Monitoring data are 

limited to determine 

if extent is more 

widespread. 

Polycyclic Aromatic Hydrocarbons

Information available 

throughout the 

Chesapeake Bay 

watershed

mostly in sediment in all 

jurisdictions including tidal

tributaries. 

limited to two areas 

(Patuxent River segments 

and DC stream segments). 

Evidence of adverse effects 

(such as liver tumors in 

fish) in several locations 

(Anacostia River and 

Elizabeth River). 

Widespread. Detections Localized. Impairments None* (see footnote)

Petroleum Hydrocarbons

Limited information for 

the Chesapeake Bay 

watershed

with areas with heavy 

boating or shipping 

activities. 

Localized. Mostly associated Localized. Impairments in 

DC and PA (based on oil and 

grease). 

Monitoring data 

are limited to 

determine if extent 

is more widespread. 

Concentration data 

are typically present in 

the form of a surrogate 

parameter “oil and 

grease”. 
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5.0 SUMMARY AND CONCLUSIONS

for pharmaceuticals. 

However, the widespread 

extent of known or 

suspected sources of 

these contaminants (e.g. 

human and animal waste 

management activities 

and infrastructure) and the 

existing occurrence data 

suggest the potential to be 

found in many locations 

throughout the watershed. 

o 

ies

f 

t 

pharmaceut

not widely available 

and represents a data 

 gap. Understanding 

the effects of individual 

and mixtures of 

pharmaceuticals, 

degradation products, 

and the adjuvants used 

in their formulations 

on resources in the Bay 

and its watershed is a 

research gap.

icals are 

data for 

Quantity of Information 

(available throughout 

the watershed or 

limited)

(localized, widespread, or 

uncertain)

widespread, or uncertain) Research Gaps

Geographic Extent Severity (localized, Monitoring and 

Pesticides

Information available 

throughout the 

Chesapeake Bay 

watershed for some 

compounds. Others 

compounds have limited 

information. 

that are in u

atrazine, simazine, 

metolachlor and their 

degradation products, 

which have been detected 

in streams and the Bay 

waters. 

Localized. Pesticides not 

currently in use, primarily 

aldrin, chlordane, dieldrin, 

DDT/DDE, heptachlor 

epoxide, mirex, and their 

degradation products. 

Detected mostly in 

sediment and fish tissues. 

Uncertain. Limited 

monitoring information 

for many other pesticides 

including some insecticides 

and fungicides. However, 

the widespread extent 

of known or suspected 

sources of these 

contaminants and the 

existing occurrence data 

suggest the potential to be 

found in many locations 

throughout the watershed. 

se, primarily 

Widespread. Herbicides 

pesticide

(see extent) still causing 

impairments at a few 

locations. 

Uncertain. Water quality 

standards not available to 

determine impairments 

for some compounds. 

Some research shows 

sublethal effects of some 

compounds (e.g. atrazine) 

at environmentally relevant 

concentrations. 

s not in use 

Localized. Organochlorine Monitoring data for many 

current use pesticides 

(e.g. some insecticides 

and fungicides) are not 

widely available and 

represent a data gap. 

Understanding the 

sub-lethal effects of 

individual (e.g., simazine 

and metolachlor) and 

mixtures of pesticides, 

degradation products, 

and the adjuvants used 

in their formulations 

on resources in the Bay 

and its watershed is a 

research gap. 

Pharmaceuticals

Limited information for 

the Chesapeake Bay 

watershed. 

monitoring information 

Uncertain Limi. ted Uncertain. Water quality 

standards not available t

determine impairments 

for these compounds. A

growing number of stud

show sublethal effects o

some compounds (e.g. 

some antidepressants, 

ciprofloxacin, 

sulfamethoxazole, and 

synthetic hormones) at 

environmentally relevan

concentrations. 

Monitoring 
Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

121

5.0 SUMMARY AND CONCLUSIONS

information for biogenic 

hormones. However, the 

widespread extent of known 

or suspected sources of these 

contaminants (human and 

animal waste management 

activities and infrastructure) 

and the existing occurrence 

data suggest the potential 

for them to be found in many 

locations throughout the 

watershed. 

Quantity of Information 

(available throughout 

the watershed or 

limited)

(localized, widespread, or 

uncertain)

widespread, or uncertain) Research Gaps

Geographic Extent Severity (localized, Monitoring and 

Household and Personal Care Products (HPCPs)

Limited information for 

the Chesapeake Bay 

watershed. 

Uncertain. Limited 

monitoring information 

for HPCPs. However, the 

widespread extent of 

known or suspected sources 

of these contaminants (e.g. 

human waste management 

activities and infrastructure) 

and the existing occurrence 

data suggest the potential 

to be found in many 

locations throughout the 

watershed. 

standards

determine impairments for 

some compounds. Some 

research shows sublethal 

effects of some compounds 

(e.g. surfactants, fragrances, 

triclosan and triclocarban) 

at environmentally relevant 

concentrations. 

 not available to 

Uncertain. Water quality Monitoring data for 

HPCPs are not widely 

available and represents a 

data gap. Understanding 

the effects of individual

and mixtures of HPCPs 

and their degradation 

products on resources in 

the Bay and its watershed 

is a research gap.

Polybrominated Diphenyl Ether Flame Retardants

Limited information for the 

Chesapeake Bay watershed. informatio

determination of extent.

n does not allow 

Uncertain. Limited monitoring 

standards not available to 

determine impairments. Some 

research shows sublethal 

effects at environmentally 

relevant concentrations.

Uncertain. Water quality Limited monitoring 

information does not allow 

determination of extent. 

Water quality standards 

do not exist and research 

is limited on effects on 

resources in the Bay and its 

watershed.

Biogenic Hormones

Limited information for the 

Chesapeake Bay watershed. 

Uncertain. Limited monitoring 

standards

determine

these compounds. Some 

research shows sublethal 

effects of biogenic hormones 

at environmentally relevant 

concentrations.

 not available to 

 impairments for 

Uncertain. Water quality Monitoring data for biogenic 

hormones are not widely 

available and represents a 

data gap. Understanding 

the effects of individual 

and mixtures of biogenic 

hormones and their 

degradation products on 

resources in the Bay and its 

watershed is a research gap.
Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

122

5.0 SUMMARY AND CONCLUSIONS

 

impairment n 

freshwater fish tissue.

Localized. Aluminum, 

chromium, iron, lead, 

manganese, zinc. Freshwater 

and sediment impairments 

in several jurisdictions and 

some exceedances of State 

standards.

s noted primarily i

. Mercury 

* “None” for research and monitoring gaps indicates that available information is sufficient to define extent and severity. Additional 

monitoring could be used to define sources and further research may be helpful to address sublethal effects on resources in the Bay and its 

watershed. 

Quantity of Information 

(available throughout 

the watershed or 

limited)

Geographic Extent 

(localized, widespread, or 

uncertain)

widespread, or uncertain) Research Gaps

Severity (localized, Monitoring and 

Metals and Metalloids

Information available 

throughout the Chesapeake 

Bay watershed 

Localized. A

chromium, iron, lead, 

manganese, zinc.

luminum, 

Widespread. Mercury Widespread None* (see footnote)
Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

123

5.0 SUMMARY AND CONCLUSIONS

Biological Effects of Toxic Contaminants on Fish and Wildlife

Additional supporting information on the toxic effects of contaminants on fish and wildlife is 

summarized to contribute to further understanding severity. This information provides insights 

for assessing the cumulative and interacting effects of toxic chemicals as well as other stressors 

on fish and wildlife.

The following indicators of compromised fish health have been observed within fish 

populations in the Chesapeake Bay watershed: increased incidence of infectious disease and 

parasite infestations contributing to increased mortality in several species of fish; feminization 

(intersex, plasma vitellogenin) of largemouth and smallmouth bass and other signs of endocrine 

disruption; reduced reproductive success and recruitment of yellow perch in certain highly 

urbanized tributaries; and tumors in bottom dwelling fish. The evidence for associations 

between toxic contaminant exposure and these indicators of compromised fish health is 

discussed. 

For wildlife in the Chesapeake Bay watershed, there are also indications of responses to 

contaminant exposure, primarily in wild birds. In a few locations, eggshell thinning associated 

with p,p’-DDE is apparent and reproduction may be impaired. There are also cases where 

organochlorine pesticide concentrations in eggs of predatory birds are at concentrations 

associated with embryo lethality. Several studies are cited in which PCB concentrations in 

addled bald eagle eggs may have been high enough to contribute to the failure to hatch. 

Detectable concentrations of PBDEs have been found in eggs from predatory birds that 

approach the lowest-observed-adverse-effect level for pipping and hatching success. 

Considerations for the Development of Federal-State 

Toxic Reduction Goals and Strategies

The findings in this report will be used during 2013 by the CBP partnership as an information 

resource to inform decisions regarding whether to establish new goals for reducing 

concentrations of toxic contaminants. If goals are established, this information resource can be 

used to inform decisions related to development of strategies by 2015 to achieve the goals. 

Considerations for development of goals and reduction strategies could also include: 

Sources – Individual contaminants have unique profiles with regard to the origin of the ongoing 

inputs and bioavailability in the watershed. Some contaminant groups, such as pesticides, 

have overlapping and varied sources and pathways to the environment. Developing a basic 
Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

124

5.0 SUMMARY AND CONCLUSIONS

understanding of the relative magnitude of sources as well as the nature of their environmental 
Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

125

5.0 SUMMARY AND CONCLUSIONS

Summary of Monitoring and Research Gaps

Monitoring gaps were identified for the following contaminant groups: dioxins and furans; 

petroleum hydrocarbons, some pesticides currently in use (e.g. insecticides and fungicides), 

pharmaceuticals, household and personal care products, flame retardants, and biogenic 

hormones. Biological monitoring at many levels of biological organization (sub-organisms to 

populations) along with systematic evaluation of water contaminants and other stressors would 

allow for more effective documentation of extent and severity.

Research that accounts for the complexities of mixture and multiple stressor effects, sublethal 

effects, nonlinear dose response curves, the role of contaminant exposures in immune response 

and subsequent pathogenic disease, would better define relations between contaminant 

exposures and potential effects in fish and wildlife. In addition, research gaps were identified 

that limit understandings of the relations between sources of these contaminants, their 

pathways to the environment, and exposures to receptor organisms. 
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APPENDIX A: SUPPLEMENTARY TABLES

172

Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

Table A-1: Water column standards - Aquatic life protection (μg/L)

Jurisdiction Partner

Parameters DC DE MD NY PA VA WV

PCBs

Polychlorinated Biphenyls PCBs FC: 0.014 FC: 0.014

SC: 0.03

FC: 0.014

SC: 0.03

FC: 0.014 FC: 0.014

SC: 0.03

FC: 0.014

PAHs

Acenaphthene FC:50 FA: 83

FC: 17

Benzidine FC: 250 FA: 0.1

FC:0.1

FA: 300

FC: 59

Benzo(a)Anthracene FA. .05

FC: .01

2-Chloronaphthalene FC: 1600 

Fluoranthene FC: 400 FA: 200

FC: 43

Naphthalene FC: 600

FA: 1.1

FC: 140

FA: 43

Phenanthrene FA: 5

FC: 1

Pesticides and Chlorinated Compounds

4,4'-DDD FA: 1.1

FC: 0.001

FA: 1.1

FC: 0.001

4,4'-DDE FA: 1.1

FC: 0.001

FA: 1.1

FC: 0.001

4,4'-DDT FA: 1.1

FC: 0.001

FA: 1.1

FC: 0.001

SA: 0.13

SC: 0.001

FA: 1.1

FC: 0.001

FA: 1.1

FC: 0.001

SA: 0.13

SC: 0.001

FA: 1.1

FC: 0.001

DDT and metabolites FA: 1.1

FC: 0.001

SA: 0.13

SC: 0.001

Aldrin FA: 3.0

FC: .04

FA: 3.0

SA: 1.3

FA: 3.0

SA: 1.3

FA: 3.0

SA: 0.1

FA: 3.0

SA: 1.3

FA: 0.003

alpha-BHC SC: 0.026

alpha-Endosulfan FA: 0.22

FC: 0.056

FA: 0.22

FC: 0.056

SA: 0.034

SC: 0.0087

FA: 0.22

FC: 0.056

FA: 0.22

FC: 0.056

SA: 0.034

SC: 0.0087

beta-BHC FA: 0.22

FC: 0.056

beta-Endosulfan FA: 0.22

FC: 0.056

FA: 0.22

FC: 0.056

SA: 0.034

SC: 0.0087

FA: 0.22

FC: 0.056

SA: 0.034

SC: 0.0087

Chlordane FA: 2.4

FC: 0.0043

FA: 2.4

FC: 0.0043

SA: 0.09

SC: 0.004

FA: 2.4

FC: 0.0043

SA: 0.09

SC: 0.004

FA: 2.4

FC: 0.0043

FA: 2.4

FC: 0.0043

SA: 0.09

SC: 0.004

FA: 2.4

FC: 0.0043

Chlorpyrifos FA: 0.083

FC: 0.041

SA: 0.011

SC: 0.0056

FA: 0.083

FC: 0.041

SA: 0.011

SC: 0.0056

FA: 0.083

FC: 0.041

SA: 0.011

SC: 0.0056

p-Cresol FA: 800

FC: 160

Demeton FC: 0.10

SC: 0.10

FC: 0.1 FC: 0.10

SC: 0.10
Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

173

APPENDIX A SUPPLEMENTARY TABLES

FC: 150 FC: 150

SA: 69

SC: 36

FC: 150

SA: 69

SC: 36

FC: 150 FC: 150 ^ FC: 150

SA: 69

SC: 36

FC: 150 ^

Table A-1: Water column standards - Aquatic life protection (μg/L)

Jurisdiction Partner

Parameters DC DE MD NY PA VA WV

Diazinon FC: 0.08 FC: 0.17

FC: 0.17

FC: 0.17

FC: 0.17

SA: 0.82

SC: 0.82

Dieldrin FA: 0.24

FC: 0.056

FA: 0.24

FC: 0.056

SA: 0.71

SC: 0.0019

FA: 0.24

FC: 0.056

SA: 0.71

SC: 0.0019

FA: 0.24

FC: 0.056

FA: 0.24

FC: 0.056

FA: 0.24

FC: 0.056

SA: 0.71

SC: 0.0019

FA: 0.24

FC: 0.056

Endosulfan FA: 0.22

FC: 0.56

FA: 0.22

FC: 0.56

SA: 0.034

SC: 0.0087

FC: 0.009

Endrin FA: 0.086

FC: 0.036

 FA: 0.086

FC: 0.036

SA: 0.037

SC: 0.0023

FA: 0.086

FC: 0.036

SA: 0.037

SC: 0.0023

FA: 0.086

FC: 0.036

FA: 0.086

FC: 0.036

FA: 0.086

FC: 0.036

SA: 0.037

SC: 0.0023

FA: 0.18

FC: 0.0023

Endrin Aldehyde

gamma - BHC (Lindane) FA: 0.095

FC: 0.08

FA: 0.095

FC: 0.16

FA: 0.095 FA: 0.095 FA: 0.095

FC: 0.16

FA: 2.0

FC: 0.08

Guthion FC: 0.01 FC: 0.01

SC: 0.01 

FC: 0.01

SC: 0.01

Heptachlor FA: 0.52

FC: 0.0038

 FA: 0.52

FC: 0.0038

SA: 0.053

SC: 0.0036

 FA: 0.52

FC: 0.0038

SA: 0.053

SC: 0.0036

FA: 0.52

FC: 0.0038 FC: 0.0038

SA: 0.053

SC: 0.0036

 FA: 0.52

Heptachlor Epoxide FA: 0.52

FC: 0.0038 FC: 0.0038

SA: 0.053

SC: 0.0036

FC: 0.0038 FC: 0.0038

SA: 0.053

SC: 0.0036

FC: 0.0038

 FA: 0.52 FA: 0.5 FA: 0.52 FA: 0.52

Hexachlorocyclohexane FA: 0.095

FC: 0.08

SA: 0.16

Malathion FC: 0.1 FC: 0.1 FC: 0.1

SC: 0.1

Methoxychlor FC: 0.03 FC: 0.03

SC: 0.03

FC: 0.03 FC: 0.03

SC: 0.03

FC: 0.03

Mirex FC: 0.001 FC: 0.001

SC: 0.001

FA: ZERO

FC: ZERO

Parathion FC: 0.013 FA: 0.065

FC: 0.013

FA: 0.065 FA: 0.065

FC: 0.013

Toxaphene FA: 0.73

FC: 0.0002 FC: 0.0002

SA: 0.21

SC: 0.0002

FC: 0.0002

SA: 0.21

SC: 0.0002

FC: 0.005 FC: 0.0002 FC: 0.0002

SA: 0.21

SC: 0.0002

FC: 0.0002

FA: 0.73 FA: 0.73 FA: 1.6 FA: 0.73 FA: 0.73 FA: 0.73

Tributyltin (TBT) FA: 0.46

FC: 0.072 FC: 0.072

SA: 0.42

SC: 0.0074

FC: 0.072

SA: 0.42

SC: 0.0074

FA: 0.46 FA: 0.46

Household and Personal Care Products

Nonylphenol FA: 28

FC: 6.6 FC: 6.6

SA: 7.0

SC: 1.7

FA: 28

Inorganics

Aluminum FA: 750*

FC: 87* FC: 750

FC: 100 FA: 750 FA: 750

Antimony FA: 1100

FC: 220

Arsenic FA: 340 FA: 340 FA: 340 FA: 340 FA: 340 ^ FA: 340 FA: 340 ^
Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

174

APPENDIX A SUPPLEMENTARY TABLES

FC: 120** FC: 120**

SA: 90

SC: 81

FC: 120**

SA: 90

SC: 81

FC: 83** FC: 120** FC: 120**

SA: 90

SC: 81

FC: 120**

Table A-1: Water column standards - Aquatic life protection (μg/L)

Jurisdiction Partner

* Aluminum Standard: pH 6.5-9.0

^ Arsenic III

** hardness = 100 mg/L CaCO3

^^ Mercury II

@Total Mercury

#

11 μg/L when hardness is ≤ ppm; 1,100 μg/L when hardness is > 75 ppm

FA: Freshwater Acute Standard

FC: Freshwater Chronic Standard

SA: Saltwater Acute Standard

SC: Saltwater Chronic Standard

Parameters DC DE MD NY PA VA WV

Barium FA: 21,000

FC: 4100

Beryllium FC: 11/1,100# FA: 130

Cadmium FA: 2.0**

FC: 0.25** FC: 0.25**

SA: 40

SC: 8.8

FC: 0.25**

SA: 40

SC: 8.8

FC: 2.1** FC: 0.25** FC: 1.1**

SA: 40

SC: 8.8

FC: 0.25**

FA: 2.0** FA: 2.0** FA: 3.8** FA: 2.0** FA: 3.9** FA: 2.0**

Chromium III FA: 570**

FC: 74** FC: 74** FC: 74** FC: 74** FC: 74** FC: 74** FC: 74**

FA: 570** FA: 570** FA: 570** FA: 570** FA: 570** FA: 570**

Chromium Vi FA: 16

FC: 11 FC: 11

SA: 1100

SC: 50

FC: 11

SA: 1100

SC: 50

FC: 11 FC: 11 FC: 11

FA: 16 FA: 16 FA: 16 FA: 16 FA: 16

Copper FA: 13**

FC: 9.0** FC: 9.0**

SA: 4.8

SC: 3.1

FC: 9.0

EA: 6.1

SA: 4.8

SC: 3.1

FC: 9.0** FC: 9.0** FC: 9.0**

SA: 9.3

SC: 6.0

FC: 9.0**

 FA: 13** FA: 13 FA: 13** FA: 13** FA: 13** FA: 13**

Iron FA: 1000 FC: 1000 FC: 1.5

Lead FA: 65**

FC: 2.5** FC: 2.5**

SA: 210

SC: 8.1

FC: 2.5**

SA: 210

SC: 8.1

FC 3.8** FC 2.5** FC: 14**

SA: 240

SC: 9.3

FC 2.5**

FA: 65** FA: 65** FA: 97** FA: 65** FA: 120** FA: 65**

Mercury FA: 1.4

FC: 0.077 FC: 0.077

SA: 1.8

SC: 0.94 

FC: 0.77

SA: 1.8

SC: 0.94

FC: 0.77 FC: 0.77^^

FA: 1.4 FA: 1.4 FA: 1.4 FA: 1.2^^ FA: 2.4@

Methylmercury FC: 0.012

Nickel FA: 470**

FC: 52** FC: 52**

SA: 74

SC: 8.2

FC: 52**

SA: 74

SC: 8.2

FC: 52** FC: 52** FC: 20**

SA: 74

SC: 8.2

FC: 52**

FA: 470** FA: 470** FA: 470** FA: 470** FA: 180** FA: 470**

Selenium FA: 20

FC: 5.0 FC: 5.0

SA: 290

SC: 71

FC: 5.0

SA: 290

SC: 71

FC: 5.0

SA: 290

SC: 71

 FA: 20 FA: 20 FC: 4.6 FC: 4.6 FA: 20

Silver FA: 3.2** FA: 3.2**

SA: 1.9 SA: 1.9 SA: 0.1 SA: 1.9

FA: 3.2** FA: 4.1** FA: 3.2** FC: 3.4** FA: 3.2**

Thallium FA: 8

FC: 20 FC: 13

FA: 65

Zinc FA: 120** FA: 120** FA: 120** FA: 120** FA: 120** FA: 120** FA: 120**
Technical Report on Toxic Contaminants in the Chesapeake Bay and its Watershed

175

APPENDIX A SUPPLEMENTARY TABLES

 

 

 

Freshwater Estuarine Waters

Table A-2: Sediment Screening Guidelines (ppb dry weight) (Buchman 2008)

Parameters TEC PEC ER-L ER-M 

PCBs

Polychlorinated Biphenyls PCBs 59.8 676 22.70 180

PAHs

Low MW PAHs 552 3,160

High MW PAHs 1,700 9,600

Total PAHs 1,610 22,800 4,022 44,792

Acenaphthene 16 500

Acenaphthylene 44 640

Anthracene 57.20 845 85.30 1,100

Benzo(a)Anthracene 108 1,050 261 1,600

Benzo(a)Pyrene 150 1,450 430 1,600

Chrysene 166 1,290 384 2,800

Dibenzo(a,h)Anthracene 33 63.40 260

Fluoranthene 423 2,230 600 5,100

Fluorene 77.40 536 19 540

2-methyl naphthalene 70 670

Naphthalene 176 561 160 2,100

Phenanthrene 204 1,170 240 1,500

Pyrene 195 1,520 665 2,600

Pesticides and Chlorinated Compounds

4,4'-DDD 4.88 28 2 20

4,4'-DDE 3.16 31.30 2.20 27

4,4'-DDT 4.16 62.90 1 7

DDx, total 5.28 572 1.58 46.10

Chlordane 3.24 17.60 0.50 6

Dieldrin 1.90 61.80 0.02 8

Endrin 2.22 207

gamma-BHC (Lindane) 2.37 4.99

Heptachlor Epoxide 2.47 16

Inorganic

Arsenic 9,790 33,000 8,200 70,000

Cadmium 990 4,980 1,200 9,600

Chromium (total) 43,400 111,000 81,000 370,000

Copper 31,600 149,000 34,000 270,000

Lead 35,800 128,000 6,700 218,000

Mercury 180 1,060 150 710

Nickel 22,700 48,600 20,900 51,600

Silver 1,000 3,700

Zinc 121,000.00 459,000 50,000 410,000