CHAPTER 3 

WATER 

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WATER CHAPTER CONTENTS 

3.1 INTRODUCTION .........................................................................................................................3-6

3.1.1 Overview of the Data .......................................................................................................3-7

3.1.2 Organization of This Chapter...........................................................................................3-8

3.2 WHAT ARE THE TRENDS IN EXTENT AND CONDITION OF FRESH SURFACE WATERS 

AND THEIR EFFECTS ON HUMAN HEALTH AND THE ENVIRONMENT?.....................3-11

3.2.1 Introduction....................................................................................................................3-11

3.2.2 ROE Indicators ..............................................................................................................3-12

INDICATOR: High and Low Stream Flows ...............................................................3-14

INDICATOR: Streambed Stability in Wadeable Streams...........................................3-19

INDICATOR: Nitrogen and Phosphorus in Wadeable Streams..................................3-22

INDICATOR: Nitrogen and Phosphorus in Streams in Agricultural Watersheds.......3-25

INDICATOR: Nitrogen and Phosphorus Discharge from Large Rivers .....................3-28

INDICATOR: Pesticides in Streams in Agricultural Watersheds ...............................3-32

INDICATOR: Benthic Macroinvertebrates in Wadeable Streams ..............................3-35

3.2.3 Discussion......................................................................................................................3-39

What These Indicators Say About Trends in the Extent and Condition of Fresh Surface 

Waters and Their Effects on Human Health and the Environment................................3-39

Limitations, Gaps, and Challenges ................................................................................3-40

3.3 WHAT ARE THE TRENDS IN EXTENT AND CONDITION OF GROUND WATER AND 

THEIR EFFECTS ON HUMAN HEALTH AND THE ENVIRONMENT? ..............................3-41

3.3.1 Introduction....................................................................................................................3-41

3.3.2 ROE Indicators ..............................................................................................................3-43

INDICATOR: Nitrate and Pesticides in Shallow Ground Water in Agricultural 

Watersheds.....................................................................................................................3-44

3.3.3 Discussion......................................................................................................................3-48

What This Indicator Says About Trends in the Extent and Condition of Ground Water 

and Their Effects on Human Health and the Environment............................................3-48

Limitations, Gaps, and Challenges ................................................................................3-48

3.4 WHAT ARE THE TRENDS IN EXTENT AND CONDITION OF WETLANDS AND THEIR 

EFFECTS ON HUMAN HEALTH AND THE ENVIRONMENT?...........................................3-50

3.4.1 Introduction....................................................................................................................3-50

3.4.2 ROE Indicators ..............................................................................................................3-52

INDICATOR: Wetland Extent, Change, and Sources of Change ...............................3-53

3.4.3 Discussion......................................................................................................................3-57

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What This Indicator Says About Trends in the Extent and Condition of Wetlands and 

Their Effects on Human Health and the Environment...................................................3-57

Limitations, Gaps, and Challenges ................................................................................3-57

3.5 WHAT ARE THE TRENDS IN EXTENT AND CONDITION OF COASTAL WATERS AND 

THEIR EFFECTS ON HUMAN HEALTH AND THE ENVIRONMENT? ..............................3-59

3.5.1 Introduction....................................................................................................................3-59

3.5.2 ROE Indicators ..............................................................................................................3-60

INDICATOR: Trophic State of Coastal Waters ..........................................................3-62

INDICATOR: Coastal Sediment Quality ....................................................................3-67

INDICATOR: Coastal Benthic Communities..............................................................3-71

INDICATOR: Submerged Aquatic Vegetation in the Chesapeake Bay......................3-74

INDICATOR: Hypoxia in the Gulf of Mexico and Long Island Sound......................3-77

INDICATOR: Harmful Algal Bloom Outbreaks Along the Western Florida Coastline .3-

81

3.5.3 Discussion......................................................................................................................3-84

What These Indicators Say About Trends in the Extent and Condition of Coastal Waters 

and Their Effects on Human Health and the Environment............................................3-84

Limitations, Gaps, and Challenges ................................................................................3-85

3.6 WHAT ARE THE TRENDS IN THE QUALITY OF DRINKING WATER AND THEIR 

EFFECTS ON HUMAN HEALTH? ...........................................................................................3-87

3.6.1 Introduction....................................................................................................................3-87

3.6.2 ROE Indicators ..............................................................................................................3-89

INDICATOR: Population Served by Community Water Systems with No Reported 

Violations of Health-Based Standards...........................................................................3-90

3.6.3 Discussion......................................................................................................................3-94

What This Indicator Says About Trends in the Quality of Drinking Water and Their 

Effects on Human Health...............................................................................................3-94

Limitations, Gaps, and Challenges ................................................................................3-94

3.7 WHAT ARE THE TRENDS IN THE CONDITION OF RECREATIONAL WATERS AND 

THEIR EFFECTS ON HUMAN HEALTH AND THE ENVIRONMENT? ..............................3-96

3.7.1 Introduction....................................................................................................................3-96

3.7.2 ROE Indicators ..............................................................................................................3-97

3.7.3 Discussion......................................................................................................................3-98

Limitations, Gaps, and Challenges ................................................................................3-98

3.8 WHAT ARE THE TRENDS IN THE CONDITION OF CONSUMABLE FISH AND 

SHELLFISH AND THEIR EFFECTS ON HUMAN HEALTH? ...............................................3-99

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3.8.1 Introduction....................................................................................................................3-99

3.8.2 ROE Indicators ............................................................................................................3-101

INDICATOR: Coastal Fish Tissue Contaminants .....................................................3-103

INDICATOR: Contaminants in Lake Fish Tissue.....................................................3-107

3.8.3 Discussion....................................................................................................................3-111

What These Indicators Say About Trends in the Condition of Consumable Fish and 

Shellfish and Their Effects on Human Health .............................................................3-111

Limitations, Gaps, and Challenges ..............................................................................3-111

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3.1 INTRODUCTION 

The nation’s water resources have immeasurable value. These resources encompass lakes, streams, 

ground water, coastal waters, wetlands, and other waters; their associated ecosystems; and the human uses 

they support (e.g., drinking water, recreation, and fish consumption). The extent of water resources (their 

amount and distribution) and their condition (physical, chemical, and biological attributes) are critical to 

ecosystems, human uses, and the overall function and sustainability of the hydrologic cycle. 

Because the extent and condition of water can affect human health, ecosystems, and critical 

environmental processes, protecting water resources is integral to EPA’s mission. EPA works in 

partnership with other government agencies that are also interested in the extent and condition of water 

resources, both at the federal level and at the state, local, or tribal level. 

In this chapter, EPA seeks to assess national trends in the extent and condition of water, stressors that 

influence water, and associated exposures and effects among humans and ecological systems. The ROE 

indicators in this chapter address seven fundamental questions about the state of the nation’s waters: 

• What are the trends in extent and condition of fresh surface waters? This question focuses 

on the nation’s rivers, streams, lakes, ponds, and reservoirs. 

• What are the trends in extent and condition of ground water? This question addresses 

subsurface water that occurs beneath the water table in fully saturated soils and geological 

formations. 

• What are the trends in extent and condition of wetlands? Wetlands—including swamps, 

bogs, marshes, and similar areas—are areas inundated or saturated by surface or ground water 

often enough and long enough to support a prevalence of vegetation typically adapted for life 

in saturated soil conditions. 

• What are the trends in extent and condition of coastal waters? Indicators in this report 

present data for coastal waters that are generally within 3 miles of the coastline (except the 

Hypoxia in the Gulf of Mexico and Long Island Sound indicator). 

• What are the trends in the quality of drinking water? People drink tap water, which comes 

from both public and private sources, and bottled water. Sources of drinking water can 

include both surface water (rivers, lakes, and reservoirs) and ground water. 

• What are the trends in the condition of recreational waters? This question addresses water 

used for a wide variety of purposes, such as swimming, fishing, and boating. 

• What are the trends in the condition of consumable fish and shellfish? This question focuses 

on the suitability of fish and shellfish for human consumption. 

These ROE questions are posed without regard to whether indicators are available to answer them. This 

chapter presents the indicators available to answer these questions, and also points out important gaps 

where nationally representative data are lacking. 

Each of the seven questions is addressed in a separate section of this chapter. However, all the questions 

are fundamentally connected—a fact that is highlighted throughout the chapter text and indicator 

summaries. All water is part of the global hydrologic cycle, and thus it is constantly in motion—whether 

it is a swiftly flowing stream or a slow-moving aquifer thousands of years old. A stream may empty into a 

larger river that ultimately discharges into coastal waters. An aquifer may be recharged by surface waters, 

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or feed surface waters or wetlands through springs and seeps. In each case, the extent and condition of 

one water resource can affect the extent and condition of another type. One example of this 

interdependence can be found in the movement of nutrients. Together, several of the ROE indicators track 

nutrient levels in water bodies ranging from small wadeable streams to the coastal waters. Additional 

ROE indicators describe some of the effects that may be associated with excess nutrients, such as 

eutrophication and hypoxia. 

In addition to the links within the water cycle, there are many connections between the extent and 

condition of water and other components of the environment. Air (addressed in Chapter 2), land (Chapter 

4), and water all are environmental media, and the condition of one medium can influence the condition 

of another. For example, contaminants can be transferred from air to water via deposition, or from land to 

water through runoff or leaching. 

Chapter 5, Human Health, and Chapter 6, Ecological Condition, examine the relationships between 

human life, ecosystems, and some of the environmental conditions that can affect them. Humans and 

ecosystems depend on water, so stressors that affect the extent and condition of water—like droughts, 

pathogens, and contaminants—may ultimately affect human health or ecological condition. 

3.1.1 Overview of the Data 

The indicators in this chapter reflect several different methods of collecting and analyzing data on the 

extent and condition of water resources (and in some cases, indicators employ a combination of methods). 

Some of the indicators in this chapter are based on probabilistic surveys, with sample or monitoring 

locations chosen to be representative of a large area (e.g., an EPA Region or the nation as a whole). 

Examples of probabilistic surveys include EPA’s Wadeable Streams Survey and National Coastal 

Assessment, and the U.S. Fish and Wildlife Service’s Wetlands Status and Trends Survey. Other 

indicators reflect targeted sampling or monitoring—for example, collecting water samples in an area 

prone to hypoxia in order to ascertain the extent and duration of a particular hypoxic event. In some cases, 

data are based on regulatory reporting, which may in turn reflect probabilistic or targeted sampling. For 

example, the ROE indicator on drinking water is based on review of monitoring conducted by water 

systems, with results reported by the states to EPA, as required by federal law. 

One of the challenges in assessing the extent and condition of water resources is that a single data 

collection method is rarely perfect for every combination of spatial and temporal domains. In general, 

there is an inherent tradeoff in representing trends in water resources. For example, a probabilistic survey 

may provide an accurate representation of national trends, but the resolution may be too low to 

definitively characterize the resource at a smaller scale. In some cases, results can be disaggregated to the 

scale of EPA Regions or ecoregions without losing precision. However, these indicators are generally not 

designed to inform the reader about the condition of his or her local water bodies, for example, or the 

quality of locally harvested fish. 

Likewise, it is often convenient to compare trends in terms of annual averages—particularly where it is 

not practical to collect data every day of the year. However, averaging and periodic sampling can obscure 

or overlook extreme events, such as spikes in water contaminants after a pesticide application or a large 

storm. Thus, representative extent or condition data cannot depict the full range of variations and 

extremes—some of which may be critical to ecosystems or to humans—that occur in smaller areas or on 

smaller time scales. 

This chapter presents only data that meet the ROE indicator definition and criteria (see Chapter 1, 

Introduction). Note that non-scientific indicators, such as administrative and economic indicators, are not 

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included in this definition. Thorough documentation of the indicator data sources and metadata can be 

found online at [insert url]. All ROE indicators were peer-reviewed during an independent peer review 

process (see [insert url] for more information). Readers should not infer that the indicators in this chapter 

reflect the complete state of knowledge. Many other data sources, publications, and site-specific research 

projects have contributed substantially to the current understanding of status and trends in water, but are 

not included in this report because they do not meet the ROE indicator criteria. 

3.1.2 Organization of This Chapter 

The remainder of this chapter is organized into seven sections corresponding to the seven questions that 

EPA seeks to answer about trends in water. Each section introduces the question and its importance, 

presents the ROE indicators used to help answer the question, and discusses what the indicators, taken 

together, say about the question. The ROE indicators include National Indicators as well as several 

Regional Indicators that meet the ROE definition and criteria and help to answer a question at a smaller 

geographic scale. Each section concludes by highlighting the major challenges to answering the question 

and identifying important information gaps. 

The table below shows the indicators used to answer each of the questions in this chapter and their 

location within this report. 

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1 Table 3.1.1. Water—ROE Questions and Indicators 

Question Indicator Name Section Page 

What are the trends in extent and condition of 

fresh surface waters and their effects on 

human health and the environment? 

High and Low Stream Flows (N) 

Streambed Stability in Wadeable 

Streams (N) 

Lake and Stream Acidity (N) 

Nitrogen and Phosphorus in 

Wadeable Streams (N) 

Nitrogen and Phosphorus in 

Streams in Agricultural 

Watersheds (N) 

Nitrogen and Phosphorus 

Discharge from Large Rivers 

(N) 

Pesticides in Streams in 

Agricultural Watersheds (N) 

Benthic Macroinvertebrates in 

Wadeable Streams (N) 

3.2.2

3.2.2

2.2.2

3.2.2

3.2.2

3.2.2

3.2.2

3.2.2

 3-14 

 3-19 

 2-62 

 3-22 

 3-25 

 3-28 

 3-32 

 3-35 

What are the trends in extent and condition of 

ground water and their effects on human 

health and the environment? 

Nitrate and Pesticides in Shallow 

Ground Water in Agricultural 

Areas (N) 

3.3.2 3-44 

What are the trends in extent and condition of 

wetlands and their effects on human health 

and the environment? 

Wetland Extent, Change, and 

Sources of Change (N) 

3.4.2 3-53 

What are the trends in extent and condition of 

coastal waters and their effects on human 

health and the environment? 

Wetland Extent, Change, and 

Sources of Change (N) 

Trophic State of Coastal Waters 

(N/R) 

Coastal Sediment Quality (N/R) 

Coastal Benthic Communities 

(N/R) 

Coastal Fish Tissue Contaminants 

(N/R) 

Submerged Aquatic Vegetation in 

the Chesapeake Bay (R) 

Hypoxia in the Gulf of Mexico and 

Long Island Sound (R) 

Harmful Algal Bloom Outbreaks 

Along the Western Florida 

Coastline (R) 

3.4.2

3.5.2

3.5.2

3.5.2

3.8.2

3.5.2

3.5.2

3.5.2

 3-53 

 3-62 

 3-67 

 3-71 

 3-103 

 3-74 

 3-77 

 3-81 

What are the trends in the quality of drinking 

water and their effects on human health? 

Population Served by Community 

Water Systems with No 

Reported Violations of HealthBased Standards (N/R) 

3.6.2 3-90 

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Question Indicator Name Section Page 

What are the trends in the condition of 

recreational waters and their effects on human 

health and the environment? 

No ROE indicators 

What are the trends in the condition of 

consumable fish and shellfish and their effects 

on human health? 

Coastal Fish Tissue Contaminants 

(N/R) 

Contaminants in Lake Fish Tissue 

(N) 

3.8.2

3.8.2

 3-103 

 3-107 

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N = National Indicator 

R = Regional Indicator 

N/R = National Indicator displayed at EPA Regional scale 

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3.2 WHAT ARE THE TRENDS IN EXTENT AND CONDITION OF FRESH 

SURFACE WATERS AND THEIR EFFECTS ON HUMAN HEALTH AND THE 

ENVIRONMENT? 

3.2.1 Introduction 

Though lakes, ponds, rivers, and streams hold less than one thousandth of a percent of the water on the 

planet, they serve many critical functions for the environment and for human life. These fresh surface 

waters sustain ecological systems and provide habitat for many plant and animal species. They also 

support a myriad of human uses, including drinking water, irrigation, wastewater treatment, livestock, 

industrial uses, hydropower, and recreation. Fresh surface waters also influence the extent and condition 

of other water resources, including ground water, wetlands, and coastal systems downstream. 

The extent of fresh surface waters reflects the influence and interaction of many stressors. It can be 

affected by direct withdrawal for drinking, irrigation, industrial processes, and other human use, as well 

as by the withdrawal of ground water, which replenishes many surface waters. Hydromodifications such 

as dam construction can create new impoundments and fundamentally alter stream flow. Land cover can 

affect drainage patterns (e.g., impervious pavement may encourage runoff or flooding). Weather 

patterns—e.g., the amount of precipitation, the timing of precipitation and snowmelt, and the conditions 

that determine evaporation rates—also affect the extent of fresh surface waters. Changing climate could 

also affect the extent of fresh surface water that is available. 

The condition of fresh surface waters reflects a range of characteristics. Physical characteristics include 

attributes such as temperature and clarity. Chemical characteristics include attributes such as salinity, 

nutrients, and chemical contaminants (including contaminants in sediments, which can impact water 

quality and potentially enter the aquatic food web). Biological characteristics include diseases, pathogens, 

and—in a broader sense—the status of plant and animal populations and the condition of their habitat. In 

addition to their effects on the environment, many of these characteristics can ultimately affect human 

health, mainly through drinking water, recreational activities (e.g., health effects in swimmers from 

pathogens and harmful algal blooms), or consumption of fish and shellfish. Because these three topics are 

complex and encompass many types of water bodies, each is addressed in greater detail in its own section 

of this report (see Sections 3.6, 3.7, and 3.8, respectively). 

Like extent, the condition of fresh surface waters can be influenced by a combination of natural and 

anthropogenic stressors, such as: 

• Point source pollution, including contaminants discharged directly into water bodies by 

industrial operations, as well as nutrients and contaminants in sewage. Even treated sewage 

contains nutrients that affect the chemical composition of the water. 

• Nonpoint source pollution, which largely reflects contaminants, nutrients, and excess 

sediment in runoff from urban and suburban areas (e.g., stormwater) and agricultural land. 

Other sources include recreational activities (e.g., boating and marinas) and acid mine 

drainage. Nonpoint source pollution can be influenced by land use—e.g., certain forestry 

techniques and agricultural practices that encourage runoff and erosion. Nonpoint sources 

tend to be more variable than point sources. For example, pesticide concentrations in streams 

reflect the location and timing of pesticide application. 

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• Air deposition. Acidic aerosols, heavy metals, and other airborne contaminants may be 

deposited directly on water or may wash into water bodies after deposition on land. Air 

deposition is a major source of mercury, for example. 

• Invasive species. Invasives are non-indigenous plant and animal species that can harm the 

environment, human health, or the economy.

1

 Invasive species can crowd out native species, 

and also may alter the physical and chemical condition of water bodies. 

• Natural factors. Precipitation determines the timing and amount of runoff and erosion, while 

other aspects of weather and climate influence heating, cooling, and mixing in lakes—which 

affect the movement of contaminants and the cycling of nutrients. The mineral composition 

of bedrock and sediment helps determine whether a water body may be susceptible to 

acidification. 

The condition of fresh surface waters also may be influenced by extent. Stream flow patterns influence 

contaminant and sediment loads, while changes in the shape of water bodies—e.g., eliminating deep pools 

or creating shallow impoundments—can change water temperature. The extent of surface waters also 

represents the extent of habitat—a key aspect of biological condition. Some plant and animal 

communities are sensitive to water level (e.g., riparian communities), while others may be adapted to 

particular seasonal fluctuations in flow. Stressors that affect extent may ultimately affect the condition of 

freshwater habitat—for example, hydromodifications that restrict the migration of certain fish species. 

3.2.2 ROE Indicators 

Eight ROE indicators characterize either the extent or the condition of fresh surface waters (Table 3.2.1). 

One of these indicators presents information about stream flow patterns, an aspect of surface water extent. 

The other seven indicators characterize various aspects of condition, including the physical condition of 

sediments, the condition of benthic communities, and the chemical condition of the water itself. Several 

of these indicators track concentrations of nutrients, which can impact many different types of water 

bodies if present in excess (e.g., through eutrophication). Supporting data come from several national 

monitoring programs: EPA’s Environmental Monitoring and Assessment Program (EMAP), EPA’s 

Wadeable Streams Assessment (WSA), EPA’s Temporally Integrated Monitoring of Ecosystems (TIME) 

and LTM (Long-Term Monitoring) projects, and three programs administered by the U.S. Geological 

Survey (the National Water Quality Assessment [NAWQA] program, the National Stream Quality 

Accounting Network [NASQAN], and the USGS stream gauge network). 

1

 National Invasive Species Council. 2005. Five year review of Executive Order 13112 on invasive species. 

Washington, DC: U.S. Department of the Interior. 

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Table 3.2.1. ROE Indicators of the Trends in Extent and Condition of Fresh Surface Waters and 

their Effects on Human Health and the Environment 

NATIONAL INDICATORS LOCATION 

High and Low Stream Flows 3.2.2 – p. 3-14 

Streambed Stability in Wadeable Streams 3.2.2 – p. 3-19 

Lake and Stream Acidity 2.2.2 – p. 2-62 

Nitrogen and Phosphorus in Wadeable Streams 3.2.2 – p. 3-22 

Nitrogen and Phosphorus in Streams in Agricultural Watersheds 3.2.2 – p. 3-25 

Nitrogen and Phosphorus Discharge from Large Rivers 3.2.2 – p. 3-28 

Pesticides in Streams in Agricultural Watersheds 3.2.2 – p. 3-32 

Benthic Macroinvertebrates in Wadeable Streams 3.2.2 – p. 3-35 

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INDICATOR: High and Low Stream Flows 

Flow is a critical aspect of the physical structure of stream ecosystems (Poff and Allan, 1995; Robinson et 

al., 2002). High flows shape the stream channel and clear silt and debris from the stream, and some fish 

species depend on high flows for spawning. Low flows define the smallest area available to stream biota 

during the year. In some cases, the lowest flow is no flow at all—particularly in arid and semi-arid 

regions where intermittent streams are common. Riparian vegetation and aquatic life in intermittent 

streams have evolved to complete their life histories during periods when water is available; however, 

extended periods of no flow can still impact their survival (Fisher, 1995). The timing of high and low 

flows also influences many ecological processes. Changes in flow can be caused by dams, water 

withdrawals, ground water pumping (which can alter base flow), changes in land cover (e.g., 

deforestation or urbanization), and weather and climate (Calow and Petts, 1992). 

This indicator, which combines two indicators presented by The Heinz Center (2005), reports on trends in 

two aspects of stream flow: 

• Flow magnitude and timing: This part of the indicator reports the percentage of streams or 

rivers throughout the contiguous 48 states that experienced major changes in the magnitude 

or the timing of average annual 1-day high flows or 7-day low flows in the 1970s, 1980s, or 

1990s, compared to a 20-year baseline period between 1930 and 1949. This indicator is based 

on 867 USGS stream gauging sites with 20 years of continuous discharge records during the 

baseline period and continuous records for the three decades between 1970 and 1999. 

• No-flow periods: This part of the indicator describes trends in no-flow periods in grassland 

and shrubland areas of the contiguous 48 states. These areas were selected for further analysis 

because they are largely arid or semi-arid, and therefore relatively water-stressed. This part of 

the indicator has two components. The first component reports the percentage of 

grassland/shrubland streams in which no-flow periods occurred during the 1950s, 1960s, 

1970s, 1980s, or 1990s. The second component reports the percentage of these streams in 

which the duration of no-flow periods during each of these decades represents an increase or 

decrease of more than 50 percent compared to the 50-year (1950-1999) average for that 

stream. Data were collected from USGS stream gauges in watersheds where at least 50 

percent of the land cover is considered grassland or shrubland per the National Land Cover 

Dataset (NLCD). These land cover designations generally correspond with three ecoregion 

types: Grassland/Steppe, California/Mediterranean, and Desert/Shrub (Bailey, 1995). The 

first component (percentage with no-flow periods) is based on 408 gauging sites; the second 

component (duration of no-flow periods) is based on 143 of these sites that had at least one 

no-flow day between 1950 and 1999. 

What the Data Show

More than half of the streams and rivers showed changes of 75 percent or more in their high or low flows 

or a shift of 60 days or more in the timing of their high or low flows, compared to the period 1930-1949 

(Exhibit 3-1). This percentage increased from 55 percent in the 1970s to 61 percent in the 1990s. About 

one-third of the streams showed moderate changes in flow (25-75 percent) or timing (30-60 days). Only 

10 percent of the streams and rivers had minimal alterations of flow of less than 25 percent or timing of 

fewer than 30 days, compared to the historical baseline period. 
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Exhibit 3-2 provides more detail about the nature of 

“major changes” in stream flow between the historical 

reference period (1930-1949) and the 1970s-1990s 

period of record. Notable trends include: 

• Approximately two-thirds of streams had 

major changes in the volume of low flow, 

with about one-third of streams showing 

substantially larger low flows throughout 

the period of record (panel A) and another 

one-third showing substantially smaller 

low flows (panel B). 

• In terms of high flow volume, more 

streams showed major decreases than 

major increases. In the 1970s and 1980s, 

only 12 percent of streams had 

substantially larger high-flow volumes 

than they had from 1930 to 1949, although 

this figure jumped to 31 percent in the 

1990s (panel D). In contrast, throughout 

the 1970s, 1980s, and 1990s, nearly 40 

percent of streams exhibited smaller high 

flows than they had during the reference 

period (panel E). 

• Between the reference period and the 

1970s-1990s period of record, about 30 

percent of streams showed major changes 

in the timing of low flows (panel C), and 

42 to 47 percent showed major changes in 

the timing of high flows (panel F). The 

number of streams in each of these 

categories increased somewhat between 

the 1970s and the 1990s. 

Overall, the percentage of streams and rivers in grassland and shrubland regions of the United States with 

periods of no flow decreased from 24 percent in the 1950s to 14 percent in the 1990s, with some variation 

by ecoregion (Exhibit 3-3). Among streams experiencing periods of no flow, the duration of these periods 

also decreased between the 1950s and 1990s (Exhibit 3-4). In the 1950s, 38 percent of these streams and 

rivers experienced no-flow periods that were at least 50 percent longer than their long-term average noflow periods during 1950-1999. By the 1990s, only 10 percent of streams fell into this category. The 

percentage of streams with no-flow periods at least 50 percent shorter than their long-term average 

increased from 23 percent in the 1950s to 63 percent in the 1990s, with a peak of 64 percent in the 1980s. 

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Indicator Limitations 

• The “magnitude and timing” component of this indicator compares stream flows in the 

decades from 1970 to 1999 with a baseline period, 1930-1949. Many dams and other 

waterworks had already been constructed by 1930, and this baseline period was characterized 

by low rainfall in some parts of the country. However, a similar analysis based on data from 

506 watersheds (USDA Forest Service, 2004) showed a tendency toward higher high- and 

low-flow rates in the decades of the 1940s, 1950s, and 1960s compared to the earlier period 

1879-1929. 

• The “dry periods” component of this indicator compares stream flows in the decades from 

1950 to 1999 with average stream flow over the full 50-year period. Like the baseline 

discussed above, this long-term average does not represent the “natural state” of stream flow 

because it postdates anthropogenic changes such as urbanization, construction of dams, etc. 

• Although the sites analyzed here are spread widely throughout the U.S., gauge placement by 

the USGS is not a random process. Gauges are generally placed on larger, perennial streams 

and rivers, and changes seen in these larger systems may differ from those seen in smaller 

streams and rivers. 

Data Sources 

The data presented in this indicator were originally published in Heinz Center (2005). The Heinz Center’s 

analysis was conducted by David Raff and N. LeRoy Poff, Colorado State University (Raff and Poff, 

2001; Raff et al., 2001; Raff, 2001), using stream flow data from the USGS National Water Information 

System database (USGS, 2005) (http://waterdata.usgs.gov/nwis). All data, including the 1930-1949 

reference data, can be downloaded from this database. Ecoregions are based on Bailey (1995). 

References 

Bailey, R.G. 1995. Description of the ecoregions of the United States. Second ed. Misc. Publ. No. 1391 

(rev). Washington, DC: USDA Forest Service. 

<http://www.fs.fed.us/land/ecosysmgmt/ecoreg1_home.html> 

Calow, P., and G.E. Petts, eds. 1992. The rivers handbook: hydrological and ecological principles. 

Volume 1. Oxford, United Kingdom: Blackwell Scientific. 

Fisher, S.G. 1995. Stream ecosystems of the western United States. In: Cushing, C.E., K.W. Cummings, 

and G.W. Minshall, eds. River and stream ecosystems, ecosystems of the world 22. New York, NY: 

Elsevier Press. 

The H. John Heinz III Center for Science, Economics, and the Environment. 2005. The state of the 

nation’s ecosystems: measuring the lands, waters, and living resources of the United States. New York, 

NY: Cambridge University Press, September 2002. Web update 2005: 

<http://www.heinzctr.org/ecosystems/report.html> 

Poff, N.L., and J.D. Allan. 1995. Functional organization of stream fish assemblages in relation to 

hydrologic variability. Ecology 76:606-27. 

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Raff, D. 2001. Addendum to report on hydrologic alteration of rivers and streams in predominantly 

grassland and shrubland ecosystems in support of the State of the Nation’s Ecosystems project for The H. 

John Heinz III Center. Colorado State University. 

Raff, D., and N. Poff. 2001. Final report on hydrologic alteration of rivers and streams in support of the 

State of the Nation’s Ecosystems project for The H. John Heinz III Center. Colorado State University. 

Raff, D., S. Howard, and N. Poff. 2001. Report on hydrologic alteration of rivers and streams in 

predominantly grassland and shrubland ecosystems in support of the State of the Nation’s Ecosystems 

project for The H. John Heinz III Center. Colorado State University.

Robinson, C.T., K. Tockner, and J.V. Ward. 2002. The fauna of dynamic riverine landscapes. Freshwater 

Biol. 47:661-677. 

USDA Forest Service. 2004. National report on sustainable forests—2003. Washington, DC. 

<http://www.fs.fed.us/research/sustain/> 

USGS. 2005. National water information system. Accessed 2005. <http://waterdata.usgs.gov/nwis> 

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INDICATOR: Streambed Stability in Wadeable Streams 

Streams and rivers adjust their channel shapes and particle sizes in response to the supply of water and 

sediments from their drainage areas, and this in turn can affect streambed stability. Lower-than-expected 

streambed stability is associated with excess sedimentation, which may result from inputs of fine 

sediments from erosion—including erosion caused by human activities such as agriculture, road building, 

construction, and grazing. Unstable streambeds may also be caused by increases in flood magnitude or 

frequency resulting from hydrologic alterations. Lower-than-expected streambed stability may cause 

stressful ecological conditions when, for example, excessive amounts of fine, mobile sediments fill in the 

habitat spaces between stream cobbles and boulders. When coupled with increased stormflows, unstable 

streambeds may also lead to channel incision and arroyo formation, and can negatively affect benthic 

invertebrate communities and fish spawning (Kaufmann et al., 1999). The opposite condition—an overly

stable streambed—is less common, and generally reflects a lack of small sediment particles. Overly stable 

streambeds can result from reduced sediment supplies or stream flows, or from prolonged conditions of 

high sediment transport without an increase in sediment supply. 

This indicator is based on the Relative Bed Stability (RBS), which is one measure of the interplay 

between sediment supply and transport. RBS is the ratio of the observed mean streambed particle 

diameter to the “critical diameter,” the largest particle size the stream can move as bedload during 

stormflows. The critical diameter is calculated from field measurements of the size, slope, and other 

physical characteristics of the stream channel (Kaufmann et al., 1999). A high RBS score indicates a 

coarser, more stable bed—i.e., streambed particles are generally much larger than the biggest particle the 

stream could carry during a stormflow. A low RBS score indicates a relatively unstable streambed, 

consisting of many fine particles that could be carried away by a stormflow. Expected values of RBS are 

based on the statistical distribution of values observed at reference sites that are known to be relatively 

undisturbed. RBS values that are substantially lower than the expected range are considered to be 

indicators of ecological stress. 

This indicator is based on data collected for the U.S. EPA’s Wadeable Streams Assessment (WSA). 

Wadeable streams are streams, creeks, and small rivers that are shallow enough to be sampled using 

methods that involve wading into the water. They typically include waters classified as 1st through 4th

order in the Strahler Stream Order classification system (Strahler, 1952). The WSA is based on a

probabilistic design, so the results from representative sample sites can be used to make a statistically 

valid statement about streambed stability in wadeable streams nationwide. 

Crews sampled 1,392 randomized sites throughout the U.S. using standardized methods (U.S. EPA, 

2004). Western sites were sampled between 2000 and 2004; eastern and central sites were all sampled in 

2004. Sites were sampled between mid-April and mid-November. At each site, crews measured substrate 

particle size, streambed dimensions, gradient, and stream energy dissipators (e.g., pools and woody 

debris), then used these factors to calculate the RBS. 

Because streambed characteristics vary geographically, streams were divided into nine ecoregions.2

 In

each ecoregion, a set of relatively undisturbed sites was sampled in order to determine the range of RBS 

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 For this analysis, the 48 contiguous states were divided into nine broad ecoregions. These “macro-level” 

ecoregions were defined by the WSA based on groupings of EPA Level III ecoregions (for a map of EPA Level III 

ecoregions, see http://www.epa.gov/wed/pages/ecoregions/level_iii.htm). A map of the nine WSA ecoregions will 

be available in the e-ROE. 
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values that would be expected among “least disturbed” streams. Next, the RBS for every site was 

compared to the distribution of RBS values among the ecoregion’s reference sites. If the observed RBS 

for a sample site was below the 5th or the 10th percentile of the regional reference distribution (depending 

on the ecoregion), the site was classified as “most disturbed.” This threshold was used because it offers a 

high degree of confidence that the observed condition is statistically different from the “least disturbed” 

reference condition. Streams with an RBS above the 25th percentile of the reference range were labeled 

“least disturbed,” indicating a high probability that the site is similar to the relatively undisturbed 

reference sites. Streams falling between the 5th and 25th percentiles were classified as “moderately 

disturbed.” Note that the “least disturbed” category may include some streams with higher-than-expected 

RBS values, which represent overly stable streambeds. Because it is more difficult to determine whether 

overly stable streambeds are “natural” or result from anthropogenic factors, this indicator only measures 

the prevalence of unstable streambeds (i.e., excess sedimentation). 

What the Data Show 

Roughly 50 percent of wadeable stream miles 

are classified as “least disturbed” with respect to 

streambed condition; that is, their streambed 

stability is close to or greater than what would 

be expected (Exhibit 3-5). Conversely, 25 

percent of the nation’s wadeable streambeds are 

significantly less stable than regional reference 

conditions for streambed stability (“most 

disturbed”), and an additional 20 percent are 

classified as “moderately disturbed.” 

Approximately 5 percent of the nation’s stream 

length could not be assessed because of missing 

or inadequate sample data. 

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Indicator Limitations 

• Samples were taken one time from each sampling location during the index period (April– 

November). Although the probability sampling design results in unbiased estimates for 

relative streambed stability in wadeable streams during the study period, RBS values may be 

different during other seasons and years because of variations in hydrology. 

• Trend data are unavailable because this is the first time that a survey on this broad scale has 

been conducted, and the survey design does not allow trends to be calculated within a single 

sampling period (2000-2004). These data will serve as a baseline for future surveys. 

Data Sources 

Aggregate data for this indicator were provided by EPA’s Wadeable Streams Assessment (WSA) (U.S. 

EPA, 2006b). Data from individual stream sites can be obtained from EPA’s STORET database (U.S. 

EPA, 2006a) (http://www.epa.gov/owow/streamsurvey/web_data.html). 

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References 

Kaufmann, P.R., P. Levine, E.G. Robison, C. Seeliger, and D. Peck. 1999. Quantifying physical habitat in 

wadeable streams. EPA/620/R-99/003. Washington, DC. 

<http://www.epa.gov/emap/html/pubs/docs/groupdocs/surfwatr/field/phyhab.pdf> 

Strahler, A.N. 1952. Dynamic basis of geomorphology. Geol. Soc. Am. Bull. 63:923-938. 

U.S. EPA. 2006a. Data from the Wadeable Streams Assessment. Accessed 2006. 

<http://www.epa.gov/owow/streamsurvey/web_data.html> 

U.S. EPA. 2006b. Wadeable Streams Assessment: a collaborative survey of the nation’s streams. EPA 

841-B-06-002. Washington, DC. 

<http://www.epa.gov/owow/streamsurvey/WSA_Assessment_Dec2006.pdf> 

U.S. EPA. 2004. Wadeable Streams Assessment: field operations manual. EPA/841/B-04/004. 

Washington, DC. <http://www.epa.gov/owow/monitoring/wsa/wsa_fulldocument.pdf> 

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INDICATOR: Nitrogen and Phosphorus in Wadeable Streams

Nitrogen and phosphorus are essential elements in aquatic ecosystems. Both nutrients are used by plants 

and algae for growth (U.S. EPA, 2005). Excess nutrients, however, can lead to increased algal production, 

and excess nutrients in streams can also affect lakes, larger rivers, and coastal waters downstream. In 

addition to being visually unappealing, excess algal growth can contribute to the loss of oxygen needed 

by fish and other animals, which in turn can lead to altered biological assemblages. Sources of excess 

nutrients include municipal sewage and septic tank drainfields, agricultural runoff, excess fertilizer 

application, and atmospheric deposition of nitrogen (Herlihy et al., 1998). 

This indicator measures total phosphorus and total nitrogen based on data collected for the U.S. EPA’s 

Wadeable Streams Assessment (WSA). Wadeable streams—streams, creeks, and small rivers that are 

shallow enough to be sampled using methods that involve wading into the water—represent a vital 

linkage between land and water. They typically include waters classified as 1st through 4th order in the 

Strahler Stream Order classification system (Strahler, 1952). The WSA is based on a probabilistic design, 

so the results from representative sample sites can be used to make a statistically valid statement about 

nitrogen and phosphorus concentrations in all of the nation’s wadeable streams. 

Crews sampled 1,392 randomized sites across the United States using standardized methods. Western 

sites were sampled between 2000 and 2004; eastern and central sites were sampled in 2004. All sites were 

sampled between mid-April and mid-November. At each site, a water sample was collected at mid-depth 

in the stream and analyzed following standard laboratory protocols (U.S. EPA, 2004a, 2004b). 

Because naturally occurring nutrient levels vary from one geographic area to another, streams were 

divided into nine ecoregions.3

 In each ecoregion, a set of relatively undisturbed sites was sampled in order 

to determine the range of nutrient concentrations that would be considered “low.” Next, observed nitrogen 

and phosphorus concentrations from all sites were compared to the distribution of concentrations among 

the ecoregion’s reference sites. If the observed result was above the 95th percentile of the ecoregion’s

reference distribution, the concentration was labeled “high.” This threshold was used because it offers a 

high degree of confidence that the observed condition is statistically different from the condition of the 

reference streams. Concentrations below the 75th percentile of the reference range were labeled “low,” 

indicating a high probability that the site is similar to the relatively undisturbed reference sites. 

Concentrations falling between the 75th and 95th percentiles were labeled “moderate.” 

3

 For this analysis, the 48 contiguous states were divided into nine broad ecoregions. These “macro-level” 

ecoregions were defined by the WSA based on groupings of EPA Level III ecoregions (for a map of EPA Level III 

ecoregions, see http://www.epa.gov/wed/pages/ecoregions/level_iii.htm). A map of the nine WSA ecoregions will 

be available in the e-ROE. 
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What the Data Show 

Nationwide, 43.3 percent of wadeable stream miles 

had low total nitrogen concentrations, while high 

nitrogen concentrations were found in 31.8 percent of 

stream miles (Exhibit 3-6). The results for total 

phosphorus are similar to those for nitrogen, with low 

concentrations in 48.8 percent of stream miles and 

high concentrations in 30.9 percent (Exhibit 3-6). The 

concentrations associated with the regional thresholds 

vary because of natural differences among the 

ecoregions. Approximately 4 percent of the nation’s 

wadeable stream length could not be assessed because 

of missing or inadequate sample data. 

Indicator Limitations 

• Samples were taken one time from each 

sampling location during the index period 

(April–November). Although the 

probability sampling design results in an 

unbiased estimate for total N and P 

concentrations in wadeable streams during 

the study period, concentrations may be 

different during other seasons. 

• Trend data are unavailable because this is the first time that a survey on this broad scale has 

been conducted, and the survey design does not allow trends to be calculated within a single 

sampling period (2000-2004). These data will serve as a baseline for future surveys. 

• Not all forms of nitrogen and phosphorus are equally bioavailable, and the ratio of nitrogen 

and phosphorus can affect the biomass and type of species of algae in streams. The forms of 

N and P and the N/P ratios may vary somewhat between the regional reference sites and the 

WSA streams. 

Data Sources 

Aggregate data for this indicator were provided by EPA’s Wadeable Streams Assessment (WSA) (U.S. 

EPA, 2006b). Data from individual stream sites can be obtained from EPA’s STORET database (U.S. 

EPA, 2006a) (http://www.epa.gov/owow/streamsurvey/web_data.html). 

References 

Herlihy, A.T., J.L. Stoddard, and C.B. Johnson. 1998. The relationship between stream chemistry and 

watershed land use data in the Mid-Atlantic region. US Water Air Soil Pollut. 105:377-386. 

Strahler, A.N. 1952. Dynamic basis of geomorphology. Geol. Soc. Am. Bull. 63:923-938. 

U.S. EPA. 2006a. Data from the Wadeable Streams Assessment. Accessed 2006. 

<http://www.epa.gov/owow/streamsurvey/web_data.html> 

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U.S. EPA. 2006b. Wadeable Streams Assessment: a collaborative survey of the nation’s streams. EPA 

841-B-06-002. Washington, DC. 

<http://www.epa.gov/owow/streamsurvey/WSA_Assessment_Dec2006.pdf> 

U.S. EPA. 2005. National estuary program—challenges facing our estuaries. Key management issues: 

nutrient overlaoding. <www.epa.gov/owow/estuaries/about3.htm> 

U.S. EPA. 2004a. Wadeable Streams Assessment: field operations manual. EPA/841/B-04/004. 

Washington, DC. <http://www.epa.gov/owow/monitoring/wsa/wsa_fulldocument.pdf> 

U.S. EPA. 2004b. Wadeable Streams Assessment: water chemistry laboratory manual. EPA/841/B04/008. Washington, DC. <http://www.epa.gov/owow/monitoring/wsa/WRS_lab_manual.pdf> 

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1 INDICATOR: Nitrogen and Phosphorus in Streams in Agricultural Watersheds 

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Nitrogen is a critical nutrient that is generally used and reused by plants within natural ecosystems, with 

minimal “leakage” into surface or ground water, where nitrogen concentrations remain very low 

(Vitousek et al., 2002). When nitrogen is applied to the land in amounts greater than can be incorporated 

into crops or lost to the atmosphere through volatilization or denitrification, however, nitrogen 

concentrations in streams can increase. The major sources of excess nitrogen in predominantly 

agricultural watersheds are fertilizer and animal waste; other sources include septic systems and 

atmospheric deposition. The total nitrogen concentration in streams is comprised of the most common 

bioavailable form (nitrate), organic nitrogen which is generally less available to biota, and nitrite and 

ammonium compounds which are typically not present in streams except in highly polluted situations. 

Excess nitrate is not toxic to aquatic life, but increased nitrogen may result in overgrowth of algae which 

can decrease the dissolved oxygen content of the water, thereby harming or killing fish and other aquatic 

species (U.S. EPA, 2005). Excess nitrogen also can lead to problems in downstream coastal waters, as 

discussed further in the N and P Discharge from Large Rivers indicator (p. 3-28). High concentrations of 

nitrate in drinking water may pose a risk of methemoglobinemia, a condition that interferes with oxygen 

transport in the blood of infants (U.S. EPA, 2004). 

Phosphorus also is an essential nutrient for all life forms, but at high concentrations the most biologically 

active form of phosphorus (orthophosphate) can cause water quality problems by overstimulating the 

growth of algae. In addition to being visually unappealing and causing tastes and odors in water supplies, 

excess algal growth can contribute to the loss of oxygen needed by fish and other animals. Elevated levels 

of phosphorus in streams can result from fertilizer use, animal wastes and wastewater, and the use of 

phosphate detergents. The fraction of total phosphorus not in the orthophosphate form consists of organic 

and mineral phosphorus fractions whose bioavailability varies widely. 

This indicator reports nitrogen and phosphorus concentrations in stream water samples collected from 

1992 to 2001 by the U.S. Geological Survey’s National Water Quality Assessment (NAWQA) program, 

which surveys the condition of streams and aquifers in study units throughout the contiguous United 

States. Specifically, this indicator reflects the condition of streams draining 111 watersheds where 

agriculture is the predominant land use, according to criteria outlined in Mueller and Spahr (2005), and 

where data are available to characterize all four species of interest (nitrate, total nitrogen, orthophosphate, 

and total phosphorus). These 111 watersheds are located in 38 of the 51 NAWQA study units (i.e., major 

river basins). Sites were chosen to avoid large point sources of nutrients (e.g., wastewater treatment 

plants). At each stream site, samples were collected 12 to 25 times each year over a 1-to-3-year period; 

this indicator is based on a flow-weighted annual average of those samples. Related indicators report the 

concentrations of nitrogen and phosphorus in small wadeable streams, regardless of land use (p. 3-22), 

and nitrate concentrations in ground water in agricultural watersheds (p. 3-44). 

For nitrogen, the indicator reports the percentage of streams with average concentrations of nitrate and 

total nitrogen in one of five ranges: less than 1 milligram per liter (mg/L); 1-2 mg/L; 2-6 mg/L; 6-10 

mg/L; and 10 mg/L or more. (This indicator measures nitrate (as N), i.e., the fraction of the material that 

is actually nitrogen.) The highest level (10 mg/L as N) represents the Maximum Contaminant Level 

(MCL) for nitrate allowed in finished drinking water in the United States (U.S. EPA, 2006). Because 

people are unlikely to drink untreated stream water, this concentration should be viewed as a point of 

reference, and not necessarily as a health risk to consumers. There is no human health guideline for total 

nitrogen and no comparable aquatic health guideline for either nitrate or total nitrogen because neither 

form represents a direct threat to organisms living in the stream. 
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Concentrations of total phosphorus and 

orthophosphate (as P) are reported in four ranges: less 

than 0.1 mg/L, 0.1-0.3 mg/L, 0.3-0.5 mg/L, and 0.5 

mg/L or more. There is currently no national water 

quality criterion for either form to protect surface 

waters because the effects of phosphorus vary by

region and are dependent on physical factors such as 

the size, hydrology, and depth of rivers and lakes. 

What the Data Show 

Average flow-weighted nitrate concentrations were 

above 2 mg/L in about half of the stream sites in these 

predominantly agricultural watersheds (Exhibit 3-7). 

About 12 percent of stream sites had nitrate 

concentrations above the federal drinking water MCL 

of 10 mg/L (the slightly smaller percentage of streams 

with total N above 10 mg/L is an artifact of the flowweighting algorithm). Nearly half of the streams 

sampled had total nitrogen concentrations in the 2-6 

mg/L range, and 72 percent had concentrations above 

2 mg/L. 

Almost 60 percent of the streams in agricultural 

watersheds had flow-weighted concentrations of 

orthophosphate (as P) of less than 0.1 mg/L. More than 

three-fourths of the streams had average annual flowweighted concentrations of total phosphorus above 0.1 

mg/L, while nearly 15 percent had total phosphorus 

concentrations above 0.5 mg/L (Exhibit 3-8). 

Indicator Limitations 

• These data represent streams draining 

agricultural watersheds in 38 of the major 

river basins (study units) sampled by the 

NAWQA program in the contiguous U.S. 

While they were chosen to be 

representative of agricultural watersheds 

across the United States, they are the 

result of a targeted sample design, and 

may not be an accurate reflection of the 

distribution of concentrations in all 

streams in agricultural watersheds in the 

U.S. 

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• This indicator does not provide information about trends over time, as the NAWQA program 

has completed only one full sampling cycle to date. Completion of the next round of 

sampling will allow trend analysis, using the data presented here as a baseline. 

• Drinking water treatment can significantly reduce concentrations of nitrate, so the levels of 

contaminants reported in this indicator are not necessarily representative of the exposures to 

people when these waters are used as public drinking water supplies. 

Data Sources 

Summary data for this indicator were provided by the U.S. Geological Survey’s National Water Quality 

Assessment (NAWQA) program. These data have been published in Mueller and Spahr (2005), along 

with the individual sampling results on which the analysis is based. 

References 

Mueller, D.K., and N.E. Spahr. 2005. Water-quality, streamflow, and ancillary data for nutrients in 

streams and rivers across the nation, 1992-2001: U.S. Geological Survey data series 152. 

<http://pubs.usgs.gov/ds/2005/152/> 

U.S. EPA. 2005. National estuary program—challenges facing our estuaries. Key management issues: 

nutrient overloading. <http://www.epa.gov/owow/estuaries/about3.htm> 

U.S.EPA. 2006. List of drinking water contaminants and MCLs. 

<http://www.epa.gov/safewater/mcl.html> 

U.S. EPA. 2004. Consumer factsheet on nitrates/nitrites. 

<http://www.epa.gov/safewater/contaminants/dw_contamfs/nitrates.html> 

Vitousek, P., H. Mooney, L. Olander, and S. Allison. 2002. Nitrogen and nature. Ambio 31:97-101. 

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1 INDICATOR: Nitrogen and Phosphorus Discharge from Large Rivers 

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Nitrogen is a critical nutrient for plants and animals, and terrestrial ecosystems and headwater streams 

have a considerable ability to capture nitrogen or to reduce it to N2 gas though the process of

denitrification. Nitrogen cycling and retention is thus one of the most important functions of ecosystems 

(Vitousek et al., 2002). When loads of nitrogen from fertilizer, septic tanks, and atmospheric deposition 

exceed the capacity of terrestrial systems (including croplands), the excess may enter surface waters, 

where it may have “cascading” harmful effects as it moves downstream to coastal ecosystems (Galloway 

and Cowling, 2002). Other sources of excess nitrogen include direct discharges from storm water or 

treated wastewater. This indicator specifically focuses on nitrate, which is one of the most bioavailable 

forms of nitrogen in bodies of water. 

Phosphorus is a critical nutrient for all forms of life, but like nitrogen, phosphorus that enters the 

environment from anthropogenic sources may exceed the needs and capacity of the terrestrial ecosystem. 

As a result, excess phosphorus may enter lakes and streams. Because phosphorus is often the limiting 

nutrient in these bodies of water, an excess may contribute to unsightly algal blooms, which cause taste 

and odor problems and deplete oxygen needed by fish and other aquatic species. In some cases, excess 

phosphorus can combine with excess nitrogen to exacerbate algal blooms (i.e., in situations where algal 

growth is co-limited by both nutrients), although excess nitrogen usually has a larger effect downstream 

in coastal waters. The most common sources of phosphorus in rivers are fertilizer and wastewater, 

including storm water and treated wastewater discharged directly into the river. In most watersheds, the 

atmosphere is not an important source or sink for phosphorus. 

This indicator tracks trends in the discharge of nitrate and phosphorus from four of the largest rivers in the 

United States: the Mississippi, Columbia, St. Lawrence, and Susquehanna. While not inclusive of the 

entire nation, these four rivers account for approximately 55 percent of all freshwater flow entering the 

ocean from the contiguous 48 states, and are geographically distributed. This indicator relies on stream 

flow and water-quality data collected by the U.S. Geological Survey (USGS), which has monitored 

nutrient export from the Mississippi River since the mid-1950s and from the Susquehanna, St. Lawrence, 

and Columbia Rivers since the 1970s. Data were collected near the mouth of each river except the St. 

Lawrence, which was sampled near the point where it leaves the United States. 

At the sites for which data are included in this indicator, USGS recorded daily stream levels and 

volumetric discharge using permanent stream gauges. Water quality samples were collected at least 

quarterly over the period of interest, in some cases up to 15 times per year. USGS calculated annual 

nitrogen load from these data using regression models relating nitrogen concentration to discharge, dayof-year (to capture seasonal effects), and time (to capture any trend over the period). These models were 

used to make daily estimates of concentrations, which were multiplied by the daily flow to calculate the 

daily nutrient load (Heinz Center, 2005). Because data on forms of nitrogen other than nitrate and nitrite 

are not as prevalent in the historical record, this indicator only uses measurements of nitrate plus nitrite. 

As nitrite concentrations are typically very small relative to nitrate, this mixture is simply referred to as 

nitrate.

What the Data Show

The Mississippi River, which drains more than 40 percent of the area of the lower 48 states, carries 

roughly 15 times more nitrate than any other U.S. river. Nitrate discharge from the Mississippi increased 

noticeably over much of the last half-century, rising from 200,000–500,000 tons per year in the 1950s and 

1960s to an average of about 1,000,000 tons per year during the 1980s and 1990s (Exhibit 3-9). Large 
year-to-year fluctuations are also evident. The 

Mississippi drains the agricultural center of the nation 

and contains a large percentage of the growing 

population, so it may not be surprising that the 

watershed has not been able to assimilate all the 

nitrogen from sources such as crop and lawn 

applications, animal manure and human wastes, and 

atmospheric deposition (e.g., Rabalais and Turner, 

2001). 

The nitrate load in the Columbia River increased to 

almost twice its historical loads during the later half 

of the 1990s, but by the last year of record (2002), the 

amount of nitrate discharged had returned to levels 

similar to those seen in the late 1970s (Exhibit 3-9). 

The St. Lawrence River showed an overall upward 

trend in nitrate discharge over the period of record, 

while the Susquehanna does not appear to have shown 

an appreciable trend in either direction. Over the 

period of record, the Columbia and St. Lawrence both 

carried an average of about 66,000 tons of nitrate per 

year, while the Susquehanna averaged 46,000 tons. 

By comparison, the Mississippi carried an average of 

770,000 tons per year over its period of record. 

The amount of phosphorus discharged decreased in 

the St. Lawrence and Susquehanna Rivers over the 

period of record (Exhibit 3-10). There is no obvious 

trend in the Mississippi and Columbia Rivers, and the 

year-to-year variability is quite large. Nitrogen and 

phosphorus discharges tend to be substantially higher 

during years of high precipitation, because of 

increased erosion and transport of the nutrients to 

stream channels (Smith et al., 2003). Over the full 

period of record, average annual phosphorus loads for 

the Mississippi, Columbia, St. Lawrence, and 

Susquehanna were 136,000; 11,000; 6,000; and 3,000 

tons, respectively. 

Indicator Limitations 

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• The indicator does not include data from numerous coastal watersheds whose human 

populations are rapidly increasing (e.g., Valigura et al., 2000). 

• It does not include smaller watersheds in geologically sensitive areas, whose ability to retain 

nitrogen might be affected by acid deposition (e.g., Evans et al., 2000). 

• It does not include forms of nitrogen other than nitrate. Although nitrate is one of the most 

bioavailable forms of nitrogen, other forms may constitute a substantial portion of the 

nitrogen load. Historically, nitrate data are more extensive than data on other forms of 

nitrogen. 

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• Not all forms of phosphorus included in 

the total phosphorus loads are equally 

capable of causing algal blooms. 

Data Sources 

A previous version of this analysis was published in 

the Heinz Center’s report, The State of the Nation’s 

Ecosystems (Heinz Center, 2005). Updated data were 

provided to EPA by USGS. USGS’s analysis was 

based on nutrient sampling and daily stream flow data, 

which can be obtained from USGS’s public databases 

(USGS, 2006a, 2006b). 

References 

Evans, C.D., A. Jenkins, and R.F. Wright. 2000. 

Surface water acidification in the South Pennines I. 

Current status and spatial variability. Environ. Pollut. 

109(1):11-20. 

Galloway, J., and E. Cowling. 2002. Reactive nitrogen 

and the world: 200 years of change. Ambio 31:64-71. 

The H. John Heinz III Center for Science, Economics, 

and the Environment. 2005. The state of the nation’s 

ecosystems: measuring the lands, waters, and living 

resources of the United States. New York, NY: 

Cambridge University Press, September 2002. Web 

update 2005: 

<http://www.heinzctr.org/ecosystems/report.html> 

Rabalais, N.N., and R.E. Turner, eds. 2001. Coastal 

hypoxia: consequences for living resources and 

ecosystems. Coastal and estuarine studies 58. 

Washington, DC: American Geophysical Union. 

Smith, S.V., D.P. Swaney, L. Talaue-McManus, J.D. 

Bartley, P.T. Sandhei, C.J. McLaughlin, V.C. Dupra, 

C.J. Crossland, R.W. Buddemeier, B.A. Maxwell, and 

F. Wulff. 2003. Humans, hydrology, and the 

distribution of inorganic nutrient loading to the ocean. 

BioScience 53:235-245. 

USGS. 2006a. National Stream Quality Accounting 

Network (NASQAN) data. Accessed 2006. 

<http://water.usgs.gov/nasqan/data/index.html> 

USGS. 2006b. National water information system. 

Accessed 2006. <http://waterdata.usgs.gov/nwis/> 

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Valigura, R., R. Alexander, M. Castro, T. Meyers, H. Paerl, P. Stacey, and R. Turner, eds. 2000. Nitrogen 

loading in coastal water bodies—an atmospheric perspective. Washington, DC: American Geophysical 

Union. 

Vitousek, P., H. Mooney, L. Olander, and S. Allison. 2002. Nitrogen and nature. Ambio 31:97-101. 

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INDICATOR: Pesticides in Streams in Agricultural Watersheds 

Pesticides are chemicals or biological agents that kill plant or animal pests and may include herbicides, 

insecticides, fungicides, and rodenticides. More than one billion pounds of pesticides (measured as 

pounds of active ingredient) are used in the United States each year to control weeds, insects, and other 

organisms that threaten or undermine human activities (Aspelin, 2003). About 80 percent of the total is 

used for agricultural purposes. Although pesticide use has resulted in increased crop production and other 

benefits, pesticide contamination of streams, rivers, lakes, reservoirs, coastal areas, and ground water can 

cause unintended adverse effects on aquatic life, recreation, drinking water, irrigation, and other uses. 

Water also is one of the primary pathways by which pesticides are transported from their application areas 

to other parts of the environment (USGS, 2000). 

This indicator is based on stream water samples collected between 1992 and 2001 as part of the U.S. 

Geological Survey’s National Water Quality Assessment (NAWQA) program, which surveys the 

condition of streams and aquifers in study units throughout the contiguous United States. Of the streams 

sampled for pesticides, this indicator focuses on 83 streams in watersheds where agriculture represents the 

predominant land use, according to criteria outlined in Gilliom et al. (2006). These 83 streams are located 

in 36 of the 51 NAWQA study units (i.e., major river basins). From each site, NAWQA collected 10 to 49 

stream water samples per year over a 1-to-3-year period to analyze for 75 different pesticides and 8 

pesticide degradation products, which together account for approximately 78 percent of the total 

agricultural pesticide application in the United States by weight during the study period (Gilliom et al., 

2006). This indicator reports the number of stream sites where the annual time-weighted average 

concentration of one or more of these pesticides or degradation products exceeds standards for aquatic or 

human health. A related indicator discusses pesticide concentrations in ground water in agricultural 

watersheds (p. 3-44). 

Three types of U.S. EPA standards and guidelines for drinking water were used as human health 

benchmarks for pesticide concentrations: Maximum Contaminant Levels (MCLs), Cancer Risk 

Concentrations (CRCs), and Lifetime Health Advisories (HA-Ls). All of these standards/guidelines are 

concentrations that pertain to lifetime exposure through drinking water (CRCs relate to potential 

carcinogens and HA-Ls relate to non-carcinogenic adverse health effects). Gilliom et al. (2006) provides a 

full list of the standards and guidelines used in this assessment; see also the list of MCLs at EPA (2006). 

More detail on these types of benchmarks, their derivation, and their underlying assumptions is provided 

in Nowell and Resek (1994). If a chemical had multiple benchmarks, the MCL was used if available; 

otherwise, the lower of the CRC (at 1 in 1,000,000 cancer risk) and HA-L values was selected. An 

exceedance was identified if a yearly, time-weighted mean concentration exceeded the corresponding 

standard or guideline (Gilliom et al., 2006). 

Several types of water quality benchmarks were used for aquatic life. Where available, data were 

compared with EPA’s acute and chronic ambient water-quality criteria for the protection of aquatic life 

(AWQC-ALs). The acute AWQC-AL is the highest concentration of a chemical to which an aquatic 

community can be exposed briefly without resulting in an unacceptable effect. The chronic AWQC-AL is 

the highest concentration to which an aquatic community can be exposed indefinitely without resulting in

an unacceptable effect. An exceedance was identified if a single sample exceeded the acute AWQC-AL or 

if a 4-day moving average exceeded the chronic AWQC-AL (per EPA’s definition of the chronic AWQCAL). Results were also compared with aquatic life benchmarks derived from toxicity values presented in 

registration and risk-assessment documents developed by EPA’s Office of Pesticide Programs (OPP). 

These benchmarks included acute and chronic values for fish and invertebrates, acute values for vascular 
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and nonvascular plants, and a value for aquatic community effects. An exceedance was identified if a 

single sample exceeded any acute benchmark or if the relevant moving average exceeded a chronic 

benchmark. More information about the derivation and application of aquatic life guidelines for this 

indicator can be found in Gilliom et al. (2006). 

What the Data Show 

In 57 percent of the streams sampled, at least one 

pesticide was detected at a concentration that exceeded

one or more aquatic life benchmarks (Exhibit 3-11). 

Roughly 10 percent of streams contained at least one 

pesticide at a concentration above the corresponding 

benchmark for human health. For reference, NAWQA 

data indicate that within this set of agricultural 

streams, at least one pesticide was present at detectable 

levels more than 90 percent of the time (Gilliom et al., 

2006). NAWQA data also indicate that pesticides in 

agricultural streams most often occur in mixtures (i.e., 

more than one compound is present in the sample) 

(Gilliom et al., 2006). The human health and 

environmental impacts of pesticide contamination, 

particularly when the pesticides occur as mixtures, are 

not well understood. 

Indicator Limitations 

• These data represent streams draining 

agricultural watersheds in 36 of the study 

units (major river basins) sampled by the 

NAWQA program in the contiguous 

United States. While they were chosen to 

be representative of agricultural 

watersheds across the nation, they are the 

result of a targeted sampling design, and 

may not be an accurate reflection of the 

distribution of concentrations in all 

streams in the nation’s agricultural watersheds. 

• This indicator does not provide information about trends over time, as the NAWQA program 

has completed only one full sampling cycle to date. Completion of the next round of 

sampling will allow trend analysis, using the data presented here as a baseline. 

• Drinking water treatment can significantly reduce concentrations of many pesticides, so the 

levels reported in this indicator are not necessarily representative of the exposures to people 

when these waters are used as public drinking water supplies. 

• Aquatic life benchmarks do not currently exist for 21 of the 83 pesticides and pesticide 

degradation products analyzed, while drinking water standards or guidelines (MCLs, CRCs, 

and HAs) do not exist for 36 of 83. Current standards and guidelines do not account for 

mixtures of pesticide chemicals and seasonal pulses of high concentrations. Possible pesticide 

effects on reproductive, nervous, and immune systems, as well as on chemically sensitive 

individuals, are not yet well understood. 

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health. Other indicators, such as Coastal Sediment Quality (p. 3-67), use aquatic life 

thresholds that are less protective. Thus, these indicators are not necessarily comparable to 

one another. 

• This indicator does not provide information on the degree to which pesticide concentrations 

exceed or fall below benchmarks. 

Data Sources 

Summary data for this indicator were provided by the U.S. Geological Survey’s National Water Quality 

Assessment (NAWQA) program. Portions of this analysis have also been published in Gilliom et al. 

(2006). Data from individual sample sites are available online in Appendix 6 of the same report 

(http://ca.water.usgs.gov/pnsp/pubs/circ1291/appendix6/). 

References 

Aspelin, A.L. 2003. Pesticide usage in the United States: trends during the 20th century. Raleigh, NC: 

Center for Integrated Pest Management, North Carolina State University. 

<http://www.pestmanagement.info/pesticide_history/index.pdf> 

Gilliom, R.J., J.E. Barbash, C.G. Crawford, P.A. Hamilton, J.D. Martin, N. Nakagaki, L.H. Nowell, J.C. 

Scott, P.E. Stackelberg, G.P. Thelin, and D.M. Wolock. 2006. Pesticides in the Nation’s Streams and 

Ground Water, 1992–2001: U.S. Geological Survey circular 1291. 

<http://ca.water.usgs.gov/pnsp/pubs/circ1291/> 

Nowell, L.H., and E.A. Resek. 1994. National standards and guidelines for pesticides in water, sediment, 

and aquatic organisms: application to water-quality assessments: Rev. Environ. Contam. Toxicol. 140:1-

164. 

U.S. EPA. 2006. List of drinking water contaminants and MCLs. 

<http://www.epa.gov/safewater/mcl.html> 

USGS (U.S. Geological Survey). 2000. Pesticides in stream sediment and aquatic biota. 

<http://ca.water.usgs.gov/pnsp/rep/fs09200/> 

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1 INDICATOR: Benthic Macroinvertebrates in Wadeable Streams 

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Freshwater benthic macroinvertebrate communities are composed primarily of insect larvae, mollusks, 

and worms. They are an essential link in the aquatic food web, providing food for fish and consuming 

algae and aquatic vegetation (Karr and Dudley, 1997). The presence and distribution of 

macroinvertebrates in streams can vary across geographic locations based on elevation, stream gradient, 

and substrate (Barbour et al., 1999). These organisms are sensitive to disturbances in stream chemistry 

and physical habitat, both in the stream channel and along the riparian zone, and alterations to the 

physical habitat or water chemistry of the stream can have direct and indirect impacts on their community 

structure. Because of their relatively long life cycles (approximately one year) and limited migration, 

benthic macroinvertebrates are particularly susceptible to site-specific stressors (Barbour et al., 1999). 

This indicator is based on data collected for the U.S. EPA’s Wadeable Streams Assessment (WSA). 

Wadeable streams are streams, creeks, and small rivers that are shallow enough to be sampled using 

methods that involve wading into the water. They typically include waters classified as 1st through 4th

order in the Strahler Stream Order classification system (Strahler, 1952). Between 2000 and 2004, crews 

sampled 1,392 sites throughout the contiguous United States using standardized methods (U.S. EPA, 

2004a, 2004b). Sites were sampled between mid-April and mid-November. At each site, a composite 

bottom sample was collected from eleven equally spaced transects within the sample reach. The WSA is 

based on a probabilistic design, so results from the sample sites can be used to make statistically valid 

statements about the percentage of wadeable stream miles that fall above or below reference values for 

the indicator. Benthic community condition was determined using two different approaches, each 

reflecting a distinct aspect of the indicator: an Index of Biological Integrity (IBI) and an 

Observed/Expected (O/E) predictive model. 

The IBI is an index that reduces complex information about community structure into a simple numerical 

value based on measures of taxonomic richness (number of taxa); taxonomic composition (e.g., insects vs. 

non-insects); taxonomic diversity; feeding groups (e.g., shredders, scrapers, or predators); habits (e.g., 

burrowing, clinging, or climbing taxa); and tolerance to stressors. Separate metrics were used for each of 

these categories in different ecoregions of the United States, based on their ability to best discriminate 

among streams.4

 Each metric was scaled against the 5th-95th percentiles for the streams in each region to 

create an overall IBI, whose value ranges from 0 to 100 (Stoddard et al., 2005). 

Once the overall IBI was established, a set of relatively undisturbed sites was selected in order to 

determine the range of IBI scores that would be expected among “least disturbed” sites. A separate 

reference distribution was developed for each ecoregion. Next, the IBI score for every sampled site was 

compared to the distribution of IBI scores among the ecoregion’s reference sites. If a site’s IBI score was 

below the 5th percentile of the regional reference distribution, the site was classified as “most disturbed.” 

This threshold was used because it offers a high degree of confidence that the observed condition is 

statistically different from the “least disturbed” reference condition. Streams with an IBI score above the 

25th percentile of the reference range were labeled “least disturbed,” indicating a high probability that the 

4

 For this analysis, the 48 contiguous states were divided into nine broad ecoregions. These “macro-level” 

ecoregions were defined by the WSA based on groupings of EPA Level III ecoregions (for a map of EPA Level III 

ecoregions, see http://www.epa.gov/wed/pages/ecoregions/level_iii.htm). A map of the nine WSA ecoregions will 

be available in the e-ROE. 
site is similar to the relatively undisturbed reference 

sites. Streams falling between the 5th and 25th 

percentiles were classified as “moderately disturbed.” 

The O/E predictive model compares the actual 

number of macroinvertebrate taxa observed at each 

WSA site (O) with the number expected (E) to be 

found at a site that is in minimally disturbed condition 

(Armitage, 1987). First, reference sites were divided 

into several groups based on the observed benthic 

assemblages, and the probability of observing each 

taxon in each group of sites was determined. Next, a 

multivariate model was used to characterize each 

group of reference sites in terms of their shared 

physical characteristics (variables that are largely 

unaffected by human influence, such as soil type, 

elevation, and latitude). This predictive model then 

was applied to each test site to determine which 

group(s) of reference sites it should be compared to. 

For each test site, the “expected” probability of 

observing each taxon was calculated as a weighted 

average based on the probability of observing that 

taxon in a particular group of reference sites and the 

probability that the test site is part of that particular 

group of sites, based on physical characteristics. The 

total “E” for the test site was generated by adding the 

probabilities of observing each of the individual taxa. 

The actual number of taxa collected at the site (O) 

was divided by “E” to arrive at an O/E ratio (Hawkins 

et al., 2000; Hawkins and Carlisle, 2001). An O/E of 

1.0 means the site’s taxa richness is equal to the 

average for the reference sites. Each tenth of a point 

below 1 suggests a 10 percent loss of taxa. 

What the Data Show 

Based on the IBI, slightly more than one quarter of 

wadeable stream miles nationwide (28.2 percent) 

were classified as “least disturbed” with respect to 

benthic macroinvertebrate condition, while 41.9 

percent were in the “most disturbed” category 

(Exhibit 3-12). Of the three major stream regions in 

the nation (see inset map, Exhibit 3-12), the eastern 

highlands had the lowest percentage of “least 

disturbed” stream miles (18.2 percent), while the 

western region had the highest percentage (45.1 

percent). 

Because there are no agreed-upon thresholds for the 

46 O/E model, the results are presented in 20 percent increments of taxa losses for the contiguous 48 states 

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(Exhibit 3-13). Nearly 40 percent (38.6 percent) of wadeable stream miles in this area have lost more than 

20 percent of their macroinvertebrate taxa, compared to comparable minimally disturbed reference sites, 

and 8.3 percent of stream miles have lost more than 60 percent of their macroinvertebrate taxa. 

Indicator Limitations 

• Although the probability sampling design results in unbiased estimates for the IBI and O/E in 

wadeable streams during the April–November index period, values may be different during 

other seasons. 

• Reference conditions for the IBI and O/E vary from one ecoregion to another in both number 

and quality, which limits the degree of ecoregional resolution at which this indicator can be 

calculated. 

• Because “E” is subject to both model error and sampling error, O/E values near 1.0 (above or 

below) do not necessarily imply a gain or loss of species relative to the reference conditions. 

• Trend data are unavailable because this is the first time that a survey on this broad scale has 

been conducted, and the survey design does not allow trends to be calculated within a single 

sampling period (2000-2004). These data will serve as a baseline for future surveys. 

Data Sources 

The results shown in Exhibit 3-12 were previously published in EPA’s 2006 Wadeable Streams 

Assessment (WSA) report (U.S. EPA, 2006c). The data in Exhibit 3-13 are based on frequency 

distributions provided by the WSA program (the 2006 report also presents results from the O/E analysis, 

but using different categories). Data from individual stream sites can be obtained from EPA’s STORET 

database (U.S. EPA, 2006a) (http://www.epa.gov/owow/streamsurvey/web_data.html). 

Ecoregions for the IBI metric are WSA “Mega Regions” based on groupings of EPA Level III Ecoregions 

(Omernik, 1987; U.S. EPA, 2006b). 

References 

Armitage, D. 1987. The prediction of the macroinvertebrate fauna of unpolluted running-water sites in 

Great Britain using environmental data. Freshwater Biol. 17:41-52. 

Barbour, M.T., J. Gerritson, B.D. Snyder, and J.B. Stribling. 1999. Rapid bioassessment protocols for use 

in streams and wadeable rivers: periphyton, benthic macroinvertebrates and fish. Second edition. 

EPA/841/B-99/002. Washington, DC: U.S. Environmental Protection Agency. 

Hawkins, C.P., and D.M. Carlisle. 2001. Use of predictive models for assessing the biological integrity of 

wetlands and other aquatic habitats. In: Rader, R.B., and D.P. Batzer, eds. Bioassessment and 

management of North American wetlands. New York, NY: John Wiley & Sons. pp. 59-83. 

Hawkins, C.P., R.H. Norris, J.N. Hogue, and J.W. Feminella. 2000. Development and evaluation of 

predictive models for measuring the biological integrity of streams. Ecol. Appl. 10:1456-1477. 

Karr, J.R., and D.R. Dudley. 1981. Ecological perspective on water quality goals. Environ. Manage. 5:55-

68. 

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Omernik, J.M. 1987. Ecoregions of the conterminous United States. Map (scale 1:7,500,000). Annals of 

the Association of American Geographers 77(1):118-125. 

Stoddard, J., D.V. Peck, S.G. Paulsen, J. Van Sickle, C.P. Hawkins, A.T. Herlihy, R.M. Hughes, F. 

Wright, P.R. Kaufmann, D.P. Larsen, G. Lomnicky, A.R. Olsen, S.A. Peterson, P.L. Ringold, and T.R. 

Whittier. 2005. An ecological assessment of western streams and rivers. EPA/620/R-05/005. Washington, 

DC: U.S. Environmental Protection Agency. 

<http://www.epa.gov/nheerl/arm/documents/EMAP.W.Assessment.final.pdf> 

Strahler, A.N. 1952. Dynamic basis of geomorphology. Geol. Soc. Am. Bull. 63:923-938. 

U.S. EPA. 2006a. Data from the Wadeable Streams Assessment. Accessed 2006. 

<http://www.epa.gov/owow/streamsurvey/web_data.html> 

U.S. EPA. 2006b. Level III ecoregions. Updated August 15, 2006. 

<http://www.epa.gov/wed/pages/ecoregions/level_iii.htm> 

U.S. EPA. 2006c. Wadeable Streams Assessment: a collaborative survey of the nation’s streams. EPA 

841-B-06-002. Washington, DC. 

<http://www.epa.gov/owow/streamsurvey/WSA_Assessment_Dec2006.pdf> 

U.S. EPA. 2004a. Wadeable streams assessment: benthic laboratory methods. EPA/841/B-04/007. 

Washington, DC. 

U.S. EPA. 2004b. Wadeable streams assessment: field operations manual. EPA/841/B-04/004. 

Washington, DC. 

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3.2.3 Discussion 

What These Indicators Say About Trends in the Extent and Condition of 

Fresh Surface Waters and Their Effects on Human Health and the 

Environment 

Although the indicators do not characterize the extent of all fresh surface waters, they do provide 

information about flow patterns in streams (Stream Flows indicator, p. 3-14). As this indicator shows, 

substantial shifts in the extent and timing of high and low flows can occur from one decade to the next. 

These shifts are particularly important in intermittent streams, where life forms may be quite sensitive to 

changes in patterns of flow and no flow. Although intermittent streams can be found throughout the 

country, the Streams Flows indicator focuses on those that occur in grassland and shrubland areas, many 

of which are arid or semi-arid and thus especially sensitive to water stress. As this indicator shows, flows 

have generally increased over the last few decades, at least on a nationwide basis. 

The physical condition of lakes and streams is in part a function of the interaction between sediment and 

water. As the Streambed Stability indicator (p. 3-19) shows, about one-fourth of the nation’s wadeable 

streams show significant evidence of excess fine sediments, which can diminish habitat. In some cases, 

excess sedimentation can reflect the influence of human stressors like erosion. Excess sedimentation also 

can be a symptom of broader changes in physical condition, such as hydromodifications that alter flow 

and sediment transport. 

The ROE indicators provide a mixed picture of the chemical condition of fresh surface waters. Acidity in 

lakes and streams is decreasing in some sensitive areas but holding steady in others (Lake and Stream 

Acidity indicator, p. 2-62), while excess nutrients are present in many streams, ranging from small 

wadeable streams to the nation’s largest rivers (three N and P indicators, pp. 3-22, 3-25, and 3-28). In 

agricultural areas, a large percentage of monitoring sites have at least one pesticide at levels that exceed 

guidelines for aquatic health (Pesticides in Agricultural Streams indicator, p. 3-32). These indicators 

reflect the influence of many stressors. For example, the two Agricultural Streams indicators (pp. 3-25 

and 3-32) demonstrate how chemicals applied to the land can ultimately affect surface waters. 

Conversely, efforts to reduce human stressors can result in improved water condition. For example, areas 

with declines in acidity correspond with areas of decreased acid deposition (Lake and Stream Acidity 

indicator, p. 2-62), while declining phosphorus loads in at least one river may be related to detergent bans 

and improved sewage treatment (N and P Discharge from Large Rivers indicator, p. 3-28). The indicators 

also reveal the influence of natural stressors, such as the role of precipitation in year-to-year variability in 

nutrient loads. 

The ROE indicators also provide a mixed picture of the biological condition of fresh surface waters. The 

indicators of extent and physical and chemical condition show a number of attributes that could 

potentially harm aquatic life, including substantial changes in high and low stream flows, a portion of 

streams with excess sedimentation, pesticides above aquatic life guidelines, and nutrients at levels that 

could encourage eutrophication. Benthic macroinvertebrate communities are particularly sensitive to 

these stressors, and thus the condition of these assemblages can provide information about the extent to 

which these stressors are causing measurable harm. About 40 percent of the nation’s wadeable stream 

miles exhibit a substantial loss (>20 percent) of macroinvertebrate taxa—approximately equal to the 

number of stream miles considered “most disturbed” when other metrics of benthic community condition 

are considered (Benthic Macroinvertebrates in Wadeable Streams indicator, p. 3-35). 

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Limitations, Gaps, and Challenges 

Although the ROE indicators provide valuable information about the extent and condition of fresh surface 

waters, there are a few general limitations to their ability to depict trends over space and time. For 

example, trends in condition may be tied to the location and timing of intermittent stressors (e.g., 

pesticide application), so indicators that assess national condition using samples that are spread out over 

time and space may obscure local conditions and extreme events. Some indicators are also restricted to 

specific study areas. For example, the two Agricultural Streams indicators (pp. 3-25 and 3-32) do not 

characterize non-agricultural watersheds, and the Lake and Stream Acidity indicator (p. 2-62) does not 

include localized acidification in the West. 

In addition to the challenges inherent in assessing fresh surface waters, there are also challenges in 

interpreting what the indicators say. Ecological responses to freshwater stressors are complex and may 

depend on the species that inhabit a particular area. In some cases—e.g., the three indicators from the 

Wadeable Streams Assessment—data must be adjusted to account for variations in regional reference 

conditions. It can also be difficult to link effects to specific stressors, as many indicators reflect the 

interplay of multiple human and natural factors. For example, local bedrock can contribute high levels of 

nutrients to some rivers, while precipitation variability can drive trends in nutrient discharge, potentially 

obscuring trends in anthropogenic stressors. 

There are no ROE indicators for a few key aspects of the extent and condition of fresh surface waters. The 

following information would help to better answer this question: 

• Information on the extent of different types of fresh surface waters, stressors to extent (e.g., 

water usage), and associated effects on ecological systems. 

• Nationally consistent information to characterize stressors to fresh surface water condition— 

specifically pollutant loadings from point and nonpoint sources. 

• Information on the condition of large rivers. The N and P Discharge from Large Rivers 

indicator (p. 3-28) describes nutrient loads at the mouth, but does not address conditions 

upstream. 

• Information on the condition of lakes. A nationally consistent indicator of lake trophic state 

could bring together several aspects of condition (e.g., physical, chemical, and biological 

parameters) related to eutrophication—a problem facing many of the nation’s lakes. 

• Information about toxic contaminants in freshwater sediments. Sediment contaminants can 

accumulate through the food web, and may ultimately impact the health of humans who 

consume fish and shellfish. 

• Information on the condition of fish communities, which can be affected by many different 

stressors. 

In addition, there are currently no ROE indicators that explicitly link human health effects to the extent or 

condition of fresh surface waters. As described in Chapter 1, this type of information gap largely reflects 

the difficulty of determining exact causation between stressors and effects. 

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3.3 WHAT ARE THE TRENDS IN EXTENT AND CONDITION OF GROUND 

WATER AND THEIR EFFECTS ON HUMAN HEALTH AND THE 

ENVIRONMENT? 

3.3.1 Introduction 

A large portion of the world’s fresh water resides underground, stored within cracks and pores in the rock 

that makes up Earth’s crust. The U.S. Geological Survey estimates that there are approximately 1,000,000 

cubic miles of ground water within one half-mile of the Earth’s surface—30 times the volume of all the 

world’s fresh surface waters.

5

 Many parts of the U.S. rely heavily on ground water for human uses (e.g., 

drinking, irrigation, industry, livestock), particularly areas with limited precipitation (e.g., the Southwest), 

limited surface water resources, or high demand from agriculture and growing populations (e.g., Florida). 

Half of the U.S. population (51 percent) relies on ground water for domestic uses.6 

Ecological systems also rely on ground water. For example, some wetlands and surface waters are fed by 

springs and seeps, which occur where a body of ground water—known as an aquifer—reaches the Earth’s 

surface. While the contribution of ground water to stream flow varies widely among streams, hydrologists 

estimate that the average contribution of ground water is an estimated 40 to 50 percent in small and 

medium sized streams. The ground water contribution to all stream flow in the U.S. may be as large as 40 

percent.7 

The extent of ground water refers to the amount available, typically measured in terms of volume or 

saturated thickness of an aquifer. The condition of ground water reflects a combination of physical, 

biological, and chemical attributes. Physical properties reflect patterns of flow—i.e., the volume, speed, 

and direction of ground water flow in a given location. Biologically, ground water can contain a variety of 

organisms, including bacteria, viruses, protozoans, and other pathogens. Ground water can also contain a 

variety of chemicals, which may occur naturally or as a result of human activities. Chemicals that may 

occur in ground water include nutrients, metals, radionuclides, salts, and organic compounds such as 

petroleum products, pesticides, and solvents. These chemicals may be dissolved in water or—in the case 

of insoluble organic contaminants—exist as undissolved plumes. 

Many stressors can affect the extent of ground water, including patterns of precipitation and snowmelt 

and human activities that change or redistribute the amount of ground water in an aquifer. One major way 

humans influence ground water extent is by withdrawing water for drinking, irrigation, or other uses (e.g., 

ground water extracted to lower the water table for mining operations). Other human activities can 

increase ground water levels, such as surface irrigation runoff recharging a shallow aquifer, or water 

pumped directly into the ground in order to store surface waters for future use, or to aid in oil and gas 

extraction. Human activities can affect ground water extent indirectly, too; for example, impervious paved 

surfaces may prevent precipitation from recharging ground water. In some cases, changes in ground water 

extent may be caused by a combination of these human and natural factors—e.g., droughts that require 

5

 U.S. Geological Survey. 1999. Ground water (general interest publication). Reston, VA. 

<http://capp.water.usgs.gov/GIP/gw_gip/> 

6

 Ibid 

7

 Alley, W.M, T.E. Reilly, and O.L. Franke. 1999. Sustainability of ground-water resources. Circular 1186. Denver, 

CO: U.S. Geological Survey. 

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humans to withdraw more water from the ground (e.g., for irrigation), while at the same time providing 

less precipitation for recharge. Additionally, some aquifers are more susceptible than others to changes in 

extent. For example, some deep aquifers may take thousands of years to recharge, particularly if they lie 

below highly impermeable confining layers. 

Aquifer depletion—i.e., decreased extent—can adversely affect the humans and ecosystems that directly 

or indirectly depend on ground water. Less ground water available for human or ecological use could 

result in lower lake levels or—in extreme cases—cause perennial streams to become intermittent or 

totally dry, thus harming aquatic and riparian plants and animals that depend on regular surface flows. 

Areas with a high water table may have plant communities that tap ground water directly with their roots, 

so even a slight lowering of the aquifer could affect native species—which in turn could benefit invasive 

species.8

 In addition, lower water table levels may lead to land subsidence and sinkhole formation in areas 

of heavy withdrawal, which can damage buildings, roads, and other structures and can permanently 

reduce aquifer recharge capacity by compacting the aquifer medium (soil or rock). Finally, changes in the 

ground water flow regime can lead to consequences such as salt water intrusion, in which saline ground 

water migrates into aquifers previously occupied by fresh ground water. 

Although aquifer depletion can have serious effects, the opposite, far less common problem—too much 

ground water—can also be detrimental. Too much ground water discharge to streams can cause erosion 

and can alter the balance of aquatic plant and animal species, as has been reported in association with 

some mining sites.9 

Like extent, condition is influenced by both natural sources and human activities. Some ground water has 

high levels of naturally occurring dissolved solids (salinity), or metals such as arsenic that can be present 

as a result of natural rock formations. Land use can affect the condition of ground water; for example, 

pesticides, fertilizers, and other chemicals applied to the land can leach into ground water, while waste 

from livestock and other animals can contribute contaminants such as nutrients, organic matter, and 

pathogens. Shallow and unconfined aquifers are particularly susceptible to this type of contamination. In 

addition, landfills may leach metals, solvents, and other contaminants into ground water (particularly 

older landfills that do not have liners and leachate collection systems). Mining operations can mobilize 

toxic metals, acidic compounds, and other substances that can impact the condition of ground water. 

Finally, chemical or biological contaminants may enter aquifers as a result of unintentional releases, 

including chemical spills on land, leaks from storage tanks, sewers or septic systems, and unplugged 

abandoned wells that allow a direct route of entry for contaminants. 

Stressors that affect ground water condition ultimately affect the condition of water available for drinking, 

irrigation, or other human needs. In some cases, treatment may be needed to ensure that finished drinking 

water does not pose risks to human health. Because drinking water can come from many different types of 

water bodies, and because of the many complex issues associated with treatment and regulation of 

drinking water, this topic is addressed in greater detail in its own section of this report, Section 3.6. The 

condition of ground water also can affect ecological systems. For example, many fish species depend on 

8

 Grantham, C. 1996. An assessment of ecological impacts of ground water overdraft on wetlands and riparian areas 

on the United States. EPA/813/S-96/001. Washington, DC: U.S. Environmental Protection Agency. 

9

 U.S. DOI (Department of the Interior). 2002. Hydrologic impacts of mining. Chapter 1. In: Permitting hydrology, a 

technical reference document for determination of probable hydrologic consequence (PHC) and cumulative 

hydrologic impact assessments (CHIA). Washington, DC. Accessed November 8, 2003. 

<http://www.osmre.gov/pdf/phc2.pdf> 

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cold, clear spring-fed waters for habitat or spawning grounds.10,11 In some cases, aquifers themselves may 

constitute ecosystems. For example, caves and sinkholes are home to many types of aquatic fauna, 

including invertebrates and fish adapted to life underground.12 Ground water can also affect the condition 

of other environmental media. For example, volatile ground water contaminants can potentially migrate 

into indoor air via soil vapor intrusion. 

In many ways, extent and condition are intertwined. For example, stressors that affect extent—such as 

withdrawal or injection—can also alter physical parameters of the ground water flow regime, such as 

velocity and direction of flow. These physical alterations can affect patterns of discharge to surface 

waters, as well as the movement of water and contaminants within the ground (e.g., salt water intrusion). 

3.3.2 ROE Indicators 

This report presents an indicator of ground water condition based on a nationwide survey of shallow wells 

in watersheds where agriculture is the predominant land use. The data come from the U.S. Geological 

Survey’s NAWQA study of major river basins with agricultural activities, representing a large portion of 

the nation’s land area. Agricultural land use is among the major sources of certain ground water 

contaminants such as nutrients and pesticides. 

Table 3.3.1. ROE Indicators of the Trends in Extent and Condition of Ground Water and their 

Effects on Human Health and the Environment 

NATIONAL INDICATORS LOCATION 

Nitrate and Pesticides in Shallow Ground Water in Agricultural 

Watersheds 

3.3.2 – p. 3-44 

10 Prichard, D., J. Anderson, C. Correll, J. Fogg, K. Gebhardt, R. Krapf, S. Leonard, B. Mitchell, and J. Stasts. 1998. 

Riparian area management: a user guide to assessing proper functioning condition and the supporting science for 

lotic areas. Technical reference 1737-15. Denver, CO: U.S. Department of the Interior, Bureau of Land 

Management, National Applied Resource Sciences Center. 126 pp. 

11 Boyd, M., and D. Sturdevant. 1997. The scientific basis for Oregon’s stream temperature standard: common 

questions and straight answers. Portland, OR: Oregon Department of Environmental Quality. 

12 Elliott, W.R. 1998. Conservation of the North American cave and karst biota. In: Wilkens, H., D.C. Culver, and 

W.F. Humphreys, eds. Subterranean biota. Amsterdam, The Netherlands: Elsevier (Ecosystems of the World series). 

pp. 665-689. Preprint online at <http://www.utexas.edu/depts/tnhc/.www/biospeleology/preprint.htm> 

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INDICATOR: Nitrate and Pesticides in Shallow Ground Water in Agricultural 

Watersheds

Nitrogen is a critical plant nutrient, and most nitrogen is used and reused by plants within an ecosystem 

(Vitousek et al., 2002), so in undisturbed ecosystems minimal “leakage” occurs into ground water, and 

concentrations are very low. When nitrogen fertilizers are applied in amounts greater than can be 

incorporated into crops or lost to the atmosphere, however, nitrate concentrations in ground water can 

increase. Elevated nitrogen levels in ground water also might result from disposal of animal waste or 

onsite septic systems. Nitrate contamination in shallow ground water (less than 100 feet below land 

surface) raises potential concerns for human health where untreated shallow ground water is used for 

domestic water supply. High nitrate concentrations in drinking water pose a risk for methemoglobinemia, 

a condition that interferes with oxygen transport in the blood of infants (U.S. EPA, 2004). 

More than one billion pounds of pesticides (measured as pounds of active ingredient) are used in the U.S. 

each year to control weeds, insects, and other organisms that threaten or undermine human activities 

(Aspelin, 2003). About 80 percent of the total is used for agricultural purposes. Although pesticide use 

has resulted in increased crop production and other benefits, pesticide contamination of ground water 

poses potential risks to human health if contaminated ground water is used as a drinking water source— 

especially if untreated. 

This indicator reports on the occurrence of nitrate and pesticides in shallow ground water in watersheds 

where agriculture is the primary land use, according to criteria outlined in Gilliom et al. (2006). Ground 

water samples were collected by the U.S. Geological Survey’s National Water Quality Assessment 

(NAWQA) program from 1992 to 2001. NAWQA surveyed 51 major river basins and aquifer regions 

across the contiguous United States during this period; the agricultural watersheds sampled were within 

34 of these study units. Although agriculture is more prevalent in some parts of the country than in others, 

the watersheds were chosen to reflect a broad range of hydrogeologic conditions and agricultural 

activities. Ground water samples were collected from existing household wells where possible and new 

observation wells otherwise, all targeted at the uppermost aquifer and avoiding locations where ground 

water condition could be biased by point sources (e.g., directly downgradient from a septic system). Most 

of the wells sampled ground water from less than 20 feet below the water table, indicating as directly as 

possible the influence of land use on shallow ground water quality. To the extent feasible, the wells were 

intended to sample recently recharged water. Most wells were sampled once; a few were sampled 

multiple times as part of a detailed nutrient study, and the results were averaged. Related indicators report 

concentrations of nutrients and pesticides in streams that drain agricultural watersheds (see the N and P in 

Agricultural Streams indicator, p. 3-25, and the Pesticides in Agricultural Streams indicator, p. 3-32). 

The nitrate component of this indicator represents 1,423 wells. Results are compared with the federal 

drinking water standard of 10 mg/L, which is EPA’s Maximum Contaminant Level (MCL) to prevent 

methemoglobinemia (U.S. EPA, 2006). MCLs are enforceable standards representing the highest level of 

a contaminant that is allowed in finished drinking water. MCLs take into account cost and best available 

treatment technology, but are set as close as possible to the level of the contaminant below which there is 

no known or expected risk to health, allowing for a margin of safety. 

Data on 75 pesticides and 8 pesticide degradation products were collected from 1,412 of the wells in the 

NAWQA study. These chemicals account for approximately 78 percent of the total agricultural pesticide 

application in the United States by weight during the study period (Gilliom et al., 2006). Three types of 

U.S. EPA human health-related standards and guidelines were used to evaluate pesticide data: Maximum 

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Contaminant Levels (MCLs) (as described above), 

Cancer Risk Concentrations (CRCs), and Lifetime 

Health Advisories (HA-Ls). In all three cases, the 

standard and guideline levels are concentrations 

pertaining to lifetime exposure through drinking water. 

The CRC is a guideline for potential carcinogens 

associated with a specified cancer risk of 1 in 

1,000,000, based on drinking water exposure over a 

70-year lifetime. The HA-L is an advisory guideline 

for drinking water exposure over a 70-year lifetime, 

considering non-carcinogenic adverse health effects. 

Specific standards and guidelines used for this 

indicator are listed in Gilliom et al. (2006), and 

additional information on these types of benchmarks, 

their derivation, and their underlying assumptions is 

provided in Nowell and Resek (1994). For this 

indicator, if a chemical had multiple benchmarks, the 

MCL took precedence; if no MCL was available, the 

lower of the CRC (at 1 in 1,000,000 cancer risk) and 

HA-L values was selected. An exceedance was 

identified if a yearly, time-weighted mean 

concentration exceeded the relevant standard or 

guideline (Gilliom et al., 2006). 

What the Data Show 

During the 1992-2001 period: 

• Nitrate concentrations were above 2 mg/L 

in 58 percent of wells sampled in areas 

where agriculture is the primary land use 

(Exhibit 3-14). By comparison, 

background nitrate levels in relatively 

undeveloped areas are generally expected 

to be below 1 mg/L (Nolan and Hitt, 

2002). 

• Nitrate concentrations in about 21 percent 

of the wells exceeded the federal drinking 

water standard (10 mg/L). 

• About 60 percent of wells had a least one 

detectable pesticide compound, and 9.4 

percent had detectable levels of five or 

more pesticides (Exhibit 3-15). According 

to NAWQA data, approximately 1 percent 

of the wells in agricultural watersheds had 

one or more pesticides at concentrations 

exceeding human health standards or 

guidelines (Gilliom et al., 2006). 

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Indicator Limitations 

• These data only represent conditions in agricultural watersheds within 34 of the major river 

basins and aquifer regions sampled by the NAWQA program from 1992 to 2001. While they 

were chosen to be representative of agricultural watersheds across the United States, they are 

the result of a targeted sample design. The data also are highly aggregated and should only be 

interpreted as an indication of national patterns. 

• This indicator does not provide information about trends over time, as the NAWQA program 

has completed only one full sampling cycle to date. Completion of the next round of 

sampling will allow trend analysis, using the data presented here as a baseline. 

• Drinking water standards or guidelines do not exist for 43 percent (36 of 83) of the pesticides 

and pesticide degradation products analyzed. Current standards and guidelines also do not 

account for mixtures of pesticide chemicals and seasonal pulses of high concentrations. 

Possible pesticide effects on reproductive, nervous, and immune systems, as well as on 

chemically sensitive individuals, are not yet well understood. 

• This indicator does not provide information on the extent to which pesticide concentrations 

exceed or fall below standards, nor the extent to which they exceed or fall below other 

reference points (e.g., Maximum Contaminant Level Goals [MCLGs] for drinking water). 

Data Sources 

Summary data for this indicator were provided by the U.S. Geological Survey’s National Water Quality 

Assessment (NAWQA) program. Pesticide data have also been published in Gilliom et al. (2006), with 

raw sampling data available online in Appendix 6 of the same report 

(http://ca.water.usgs.gov/pnsp/pubs/circ1291/appendix6/). Summary data for nitrate have not yet been 

published; however, data from individual sample sites can be obtained from NAWQA’s online data 

warehouse (USGS, 2006). 

References 

Aspelin, A.L. 2003. Pesticide usage in the United States: trends during the 20th century. Raleigh, NC: 

Center for Integrated Pest Management, North Carolina State University. 

<http://www.pestmanagement.info/pesticide_history/index.pdf>. 

Gilliom, R.J., J.E. Barbash, C.G. Crawford, P.A. Hamilton, J.D. Martin, N. Nakagaki, L.H. Nowell, J.C. 

Scott, P.E. Stackelberg, G.P. Thelin, and D.M. Wolock. 2006. Pesticides in the Nation’s Streams and 

Ground Water, 1992–2001: U.S. Geological Survey circular 1291. 

<http://pubs.usgs.gov/circ/2005/1291/> 

Nolan, B.T., and K.J. Hitt. 2002. Nutrients in shallow ground waters beneath relatively undeveloped areas 

in the conterminous United States: U.S. Geological Survey water resources investigation report 02-4289. 

17 pp. <http://water.usgs.gov/nawqa/nutrients/pubs/wri02-4289/wri02-4289.pdf> 

Nowell, L.H., and E.A. Resek. 1994. National standards and guidelines for pesticides in water, sediment, 

and aquatic organisms: application to water-quality assessments: Rev. Environ. Contam. Toxicol. 140:1-

164. 

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U.S. EPA. 2006. List of Drinking Water Contaminants and MCLs. 

<http://www.epa.gov/safewater/mcl.html> 

U.S. EPA. 2004. Consumer factsheet on nitrates/nitrites. 

<http://www.epa.gov/safewater/contaminants/dw_contamfs/nitrates.html> 

USGS. 2006. USGS National Water Quality Assessment data warehouse. Accessed 2006. 

<http://infotrek.er.usgs.gov/traverse/f?p=NAWQA:HOME:14620854136944137608> 

Vitousek, P., H. Mooney, L. Olander, and S. Allison. 2002. Nitrogen and nature. Ambio 31:97-101. 

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3.3.3 Discussion 

What This Indicator Says About Trends in the Extent and Condition of 

Ground Water and Their Effects on Human Health and the Environment 

The Nitrate and Pesticides in Ground Water indicator (p. 3-44) describes the extent to which the condition 

of shallow ground water may be influenced by human stressors—in this case, certain chemicals applied to 

land in agricultural areas. Collectively, the agricultural watersheds sampled across the nation had average 

nitrate concentrations that were substantially higher than the background levels one might expect in an 

undisturbed watershed. Nitrate concentrations exceeded MCLs for nitrate in one-fifth of the wells, though 

this does not necessarily reflect the condition of the water people drink if it is tested and treated. Nitrate 

concentrations were often high enough that they could impact ecological systems upon being introduced 

into surface waters.13,14 Pesticide compounds were detected frequently (more than half of the shallow 

wells sampled). However, detected pesticide concentrations rarely exceeded human health-based 

reference points in the samples collected for this indicator. 

Limitations, Gaps, and Challenges 

One challenge in answering this question is that there are currently no national indicators of ground water 

extent. Comprehensive national data do not exist, particularly in terms of real-time water level 

monitoring. Statistics on water use and withdrawal might be considered a surrogate for ground water 

extent, but because withdrawal is but one factor that affects extent (other factors include recharge rate and 

flow patterns), the relationship between withdrawal and extent differs from one location to another. Thus, 

the issue of extent currently represents an information gap. 

There are also several limitations, gaps, and challenges in addressing the issue of ground water condition. 

One notable limitation to the Nitrate and Pesticides in Ground Water indicator (p. 3-44) is that it does not 

provide information about trends over time. The indicator is also limited in its ability to represent the 

condition of entire aquifers. Because ground water condition is vertically heterogeneous, results from one 

depth do not necessarily represent other depths. This indicator characterizes the uppermost layer of 

shallow aquifers, which are used by many private wells. It does not provide information about the 

condition of deeper aquifers, which are more likely to be used for public water supplies. 

The Nitrate and Pesticides in Ground Water indicator provides a representative national picture of shallow 

ground water condition in agricultural watersheds. At the present time, similar indicators do not exist for 

ground water in watersheds with non-agricultural land uses. Non-agricultural watersheds—particularly 

urban areas—reflect a different set of stressors, and to some extent a different set of chemicals (i.e., 

VOCs and hydrocarbons like MTBE15). Because many ground water stressors in urban areas are localized 

13 Howarth, R., D. Anderson, J. Cloern, C. Elfring, C. Hopkinson, B. Lapointe, T. Malone, N. Marcus, K. 

McGlathery, A. Sharpley, and D. Walker. 2000. Nutrient pollution of coastal rivers, bays, and seas. Issues in 

ecology, number 7. Washington, DC: Ecological Society of America. 

14 Jackson, R.., S. Carpenter, C. Dahm, D. McKnight, R. Naiman, S. Postel, and S. Running. 2001. Water in a 

changing world. Issues in ecology, number 9. Washington, DC: Ecological Society of America. 

15 Delzer, G.C., and T. Ivahnenko, 2003. Occurrence and temporal variability of methyl tert-butyl ether (MTBE) and 

other volatile organic compounds in select sources of drinking water: results of the focused survey. USGS series: 

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events such as plumes resulting from chemical spills or underground storage tank (UST) leaks, they may 

be harder to characterize on a national level—a potential challenge to gathering more information about 

ground water condition. Salt water intrusion is another issue that tends to occur locally, and for which 

national-scale data are not available. 

water-resources investigations report. Report no. 2002-4084. Reston, VA: U.S. Geological Survey. 

<http://sd.water.usgs.gov/nawqa/pubs/wrir/wrir02_4084.pdf> 

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3.4 WHAT ARE THE TRENDS IN EXTENT AND CONDITION OF WETLANDS AND 

THEIR EFFECTS ON HUMAN HEALTH AND THE ENVIRONMENT? 

3.4.1 Introduction 

The United States has many types of wetlands, which include marshes, swamps, bogs, and similar marine, 

estuarine, or freshwater areas that are periodically saturated or covered by water. Wetlands are an integral 

part of the landscape because they provide habitat for a diverse array of plants and animals, act as buffers 

to flooding and erosion, and serve as key links in the global water and biogeochemical cycles. 

In terms of extent, wetlands currently cover 5.5 percent of the surface area of the contiguous 48 states, 

with freshwater wetlands accounting for nearly 95 percent of the current wetland acreage and marine and 

estuarine wetlands accounting for the remaining 5 percent.16 Condition is somewhat harder to measure, as 

it reflects a combination of physical, chemical, and biological attributes. To be in healthy condition, 

however, a wetland should generally demonstrate good water quality and support native plant and animal 

communities, without the presence of invasive non-indigenous species. A healthy wetland should not 

show signs of stress related to substantial degradation or cumulative effects of smaller degradations, and 

should be free of modifications that restrict water flow into, through, or out of the wetland, or that alter 

patterns of seasonality. 

Wetlands can be classified by many different attributes. First, they can be divided by basic location— 

freshwater, marine, or estuarine. Wetlands also may be classified based on dominant vegetation type. For 

example, swamps are dominated by trees and shrubs, while marshes are characterized by non-woody, 

emergent (vertically oriented) plants like grasses and sedges. Other characteristics used to classify 

wetlands include soil type, water source, and the length of time a given wetland is saturated. 

The structure and function of any given wetland will be governed by a combination of interrelated factors, 

including topography, underlying geology (e.g., mineral composition), the abundance and movement of 

water (hydrology), and weather and climate. These factors ultimately determine which plant and animal 

species will thrive in a given wetland. 

All wetlands share a few basic physical, chemical, and biological attributes. By definition, all wetlands 

are saturated or covered by water at least periodically, and wetland vegetation is adapted to these 

conditions. Thus, wetlands are like sponges, with a natural ability to store water. Wetlands also tend to 

have highly developed root systems that anchor trees and other vegetation in place. This web of roots not 

only holds the soil in place, but also filters pollutants out of the water as it flows through. 

Because of their physical, chemical, and biological properties, wetlands serve many important 

environmental functions. They play an important role in improving natural water quality by filtering 

pollutants. This function is particularly important to human health because it may affect the condition of 

waters used as a source of drinking water—a topic described in greater detail in Section 3.6. Wetlands 

also act as a buffer to protect the shoreline from erosion and storm damage. Because of their sponge-like 

capacity to absorb water, wetlands slow the water’s momentum and erosive potential and reduce flood 

16 Dahl, T.E. 2006. Status and trends of wetlands in the conterminous United States 1998 to 2004. Washington, DC: 

U.S. Department of the Interior, Fish and Wildlife Service. <http://wetlandsfws.er.usgs.gov/status_trends/> 

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heights. During dry periods, the “sponge” releases water, which is critical in maintaining the base flow of 

many surface water systems. 

Wetlands are also among the most biologically productive natural ecosystems in the world. Microbial 

activity in wetlands enriches the water and soil with nutrients. As the interface between terrestrial and 

aquatic ecological systems, wetlands provide food and habitat for many plant and animal species, 

including rare and endangered species. Because of these functions, wetlands support a number of human 

activities, including commercial fishing, shellfishing, and other industries, as well as recreation, 

education, and aesthetic enjoyment. 

In addition, wetlands play a role in global biogeochemical cycles, particularly those driven in part by the 

microbial processes that occur in wetlands (e.g., the mineralization of sulfur and nitrogen from decaying 

plants and the methylation of mercury). Plant growth in wetlands provides a “sink” for many chemicals 

including atmospheric carbon. If a wetland is disturbed or degraded, these cycles can be altered and some 

of the chemicals may be released. 

The extent of wetlands can be affected by a variety of natural stressors, such as erosion, land subsidence, 

changes in precipitation patterns (e.g., droughts), sea level change, hurricanes, and other types of storms. 

However, the vast majority of wetland losses and gains over the last few centuries have occurred as a 

result of human activity.17 For years, people have drained or filled wetlands for agriculture or urban and 

suburban development, causing habitat loss or fragmentation as well as a decline in many of the other 

important functions outlined above, such as improving water quality. Conversely, other human activities 

may increase the extent of wetlands—for example, creating shallow ponds or re-establishing formerly 

drained or modified wetlands on farmlands. 

Wetland extent may influence condition, as wetland loss may result in added stress to remaining 

wetlands. For example, if fewer wetlands are available to filter pollutants from surface waters, those 

pollutants could become more concentrated in remaining downgradient wetlands. Wetland loss and 

fragmentation also lead to decreases in habitat, landscape diversity, and the connectivity among aquatic 

resources (i.e., fragmented wetlands essentially become isolated wildlife refuges). Thus, stressors that 

affect extent may ultimately affect condition as well. 

Wetland condition also reflects the influence of stressors that affect topography, hydrology, climate, water 

condition, and biodiversity. For example, human modifications such as pipes and channels can alter the 

topography, elevation, or hydrology of wetlands, while withdrawal of ground water or upstream surface 

waters can directly reduce inflow. Natural forces and human activities (e.g., hurricanes, sea level change, 

and certain agricultural and forestry practices) can also affect wetlands through increased erosion or 

sedimentation. Pollutants in ground water and fresh surface waters that flow into wetlands may be toxic to 

plants and animals, and may also accumulate in wetland sediments. In addition, invasive species can alter 

the composition of wetland communities. Some of the most well known invasives in the U.S. are wetland 

species, including plants such as phragmites and purple loosestrife and animals such as the nutria (a South 

American rodent introduced to the Chesapeake and Gulf states). 

17 Dahl, T.E. 2000. Status and trends of wetlands in the conterminous United States, 1986 to 1997. Washington, DC: 

U.S. Department of the Interior, U.S. Fish and Wildlife Service. 

<http://wetlands.fws.gov/bha/SandT/SandTReport.html> 

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Another key stressor to wetlands is conversion from one wetland type to another. Although conversion 

can occur naturally through plant succession (such as marshes turning into forested wetlands over time), 

human activities can cause more drastic changes, such as clearing trees from a forested wetland, 

excavating a marsh to create an open water pond, or introducing certain invasive species (e.g., the nutria, 

which converts tidal marsh to open water by removing vegetation). Even if wetland extent is not altered, 

conversion from one type to another has a major ecological impact by altering habitat types and 

community structure. 

3.4.2 ROE Indicators 

An ROE indicator describes trends in wetland extent, as well as specific activities that have contributed to 

recent wetland losses and gains (Table 3.4.1). Data were collected as part of the U.S. Fish and Wildlife 

Service’s Wetlands Status and Trends survey, a probabilistic national survey of wetland acreage 

conducted approximately every 10 years for the past half-century. There is no ROE indicator for wetland 

condition. 

Table 3.4.1. ROE Indicators of the Trends in Extent and Condition of Wetlands and their Effects 

on Human Health and the Environment 

NATIONAL INDICATORS LOCATION 

Wetland Extent, Change, and Sources of Change 3.4.2 – p. 3-53 

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1 INDICATOR: Wetland Extent, Change, and Sources of Change 

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Wetlands support a variety of fish and wildlife species and contribute to the aesthetic and environmental 

quality of the U.S. Millions of Americans use freshwater wetlands annually for hunting, fishing, bird 

watching and other outdoor activities. Estuarine wetlands provide valuable nursery, feeding, breeding, 

staging, and resting areas for an array of fish, shellfish, mammals, and birds (Dahl, 2000). In addition, 

wetlands serve as ground water recharge areas and filter contaminants from surface runoff (Mitsch and 

Gosselink, 1986). Destruction or alteration of wetlands, therefore, can have wide-ranging biological and 

hydrological impacts. 

Various lines of evidence suggest when European settlers first arrived, wetland acreage in the area that 

would become the contiguous 48 states was more than twice what it is today (Dahl, 1990). Since then, 

extensive losses have occurred due to draining and filling. In addition to the sheer loss of wetland 

acreage, major ecological impacts also have resulted from the conversion of one wetland type to another, 

such as clearing trees from a forested wetland or excavating a shallow marsh to create an open water 

pond. These types of conversions change habitat types and community structure in watersheds and impact 

the animal communities that depend on them (Dahl, 2000). 

This indicator presents data from the U.S. Fish and Wildlife Service’s Wetlands Status and Trends 

survey. Conducted approximately every 10 years, this survey provides an estimate of the extent of all 

wetlands in the contiguous U.S., regardless of land ownership. The Status and Trends survey uses a 

probabilistic design, based initially on stratification of the 48 contiguous states by state boundaries and 35 

physiographic subdivisions. Within these subdivisions are located 4,375 randomly selected, four square 

mile (2,560 acre) sample plots. These plots are examined with the use of aerial imagery. Although the 

imagery ranges in scale and type, most are 1:40,000 scale, color infrared from the National Aerial 

Photography Program. Field verification is conducted to address questions of image interpretation, land 

use coding, and attribution of wetland gains or losses; plot delineations are also completed. In the 1980s 

to 1990s analysis, 21 percent of the sample plots were field-verified; in the most recent analysis, 32 

percent were field-verified (Dahl, 2000, 2006). The Fish and Wildlife Service used the Cowardin et al. 

(1979) definition of wetlands, which is part of the draft national standard for wetland mapping, 

monitoring, and data reporting as determined by the Federal Geographic Data Committee. 

This indicator shows trends in the total extent of wetlands, as well as the extent of several types of 

freshwater and intertidal wetlands. In this analysis, freshwater wetlands include forested, shrub, emergent, 

and non-vegetated wetlands (e.g., shallow ponds). Intertidal wetlands include marine areas (e.g., tidal flats 

and sandbars) and estuarine areas (vegetated or not) that are exposed and flooded by the tides. Data on 

wetland extent are described from several Status and Trends analyses: 1950s-1970s, 1970s-1980s, 1980s1990s, and 1998-2004 (Frayer et al., 1983; Dahl and Johnson, 1991; Dahl, 2000, 2006). For the most 

recent period, the indicator also describes sources of wetland loss or gain, which the survey divided into 

five distinct categories along with an “other” category (Dahl, 2006). 

What the Data Show

Total wetland acreage declined over the last 50 years, but the rate of loss appears to have slowed over 

time. From the 1950s to 1970s, an average of 458,000 acres was lost per year (Exhibit 3-16). By the 

1986-1997 period, the loss rate had declined to 58,600 acres per year; and in the most recent study period, 

1998-2004, wetland area increased at a rate of 32,000 acres per year (Exhibit 3-16). 
Gains and losses have varied by wetland type. 

Freshwater forested wetlands, which make up more 

than half of all freshwater wetlands, lost acreage from 

the 1950s to the 1990s but have shown gains since 

1998 (Exhibit 3-17, panel A). Freshwater emergent 

wetlands have continued to lose acreage, although the 

rate of loss has slowed recently (panel C). Among 

freshwater categories, forested wetlands have sustained 

the greatest absolute losses since the 1950s, about 9 

million acres, while emergent wetlands have shown the 

largest percentage loss (about 21 percent). Conversely, 

the extent of freshwater shrub wetlands increased until 

1998 but declined thereafter, suggesting that some of 

the gains and losses in specific categories may reflect 

conversion rather than outright wetland loss or gain 

(Dahl, 2006; Exhibit 3-17, panel B). Shallow 

freshwater ponds, meanwhile, have increased steadily 

throughout the last 50 years, with current acreage more 

than twice what it was in the 1950s (panel D). These 

wetlands account for a large percentage of the recent 

gains illustrated in Exhibit 3-17 (Dahl, 2006). 

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Since the 1950s, intertidal wetland 

acreage has decreased by about 

700,000 acres, or 12 percent (Exhibit 

3-18, panel A). This category 

includes marine, estuarine vegetated, 

and estuarine non-vegetated 

wetlands. Both estuarine types lost 

acreage overall, with estuarine 

vegetated wetlands, the predominant 

type, losing over 400,000 acres 

(panel B). Long-term trends, 

however, indicate that losses of 

intertidal wetlands have slowed over 

time, with estuarine non-vegetated 

wetlands actually gaining acreage 

from 1998 to 2004 (panel C). 

Between 1998 and 2004, urban 

development, rural development, 

conversion to deepwater, and 

silviculture all contributed to losses 

in wetland acreage (Exhibit 3-19). 

However, the net change in wetland 

acreage during this period was 

positive, due largely to wetland 

creation and restoration on 

agricultural lands (70,770 acres) and 

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on lands classified as “other” (349,600 acres). This “other” category includes conservation lands, areas in 

transition from one land use to another, and other lands that do not fall into the major land use categories 

as defined in Dahl (2006). 

Indicator Limitations 

• Different methods were used in some of 

the early schemes to classify wetland 

types. As methods and spatial resolution 

have improved over time, acreage data 

have been adjusted, resulting in changes 

in the overall wetland base over time, thus 

reducing the accuracy of the trend. 

• Ephemeral waters and effectively drained 

palustrine wetlands observed in farm 

production are not recognized as wetland 

types by the Status and Trends survey and 

are therefore not included in the indicator. 

• Forested wetlands are difficult to 

photointerpret and are generally 

underestimated by the survey. 

• The aerial imagery used for this survey 

generally does not allow detection of 

small, isolated patches of wetland less 

than about an acre. 

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• Alaska and Hawaii are not included in the Status and Trends survey. 

• This survey does not include Pacific coast estuarine wetlands such as those in San Francisco 

Bay, Puget Sound, or Coos Bay, Oregon. 

Data Sources 

Data for this indicator were obtained from Dahl (2006). Historical trends are based on data originally 

presented in earlier Fish and Wildlife Service reports (Dahl, 2000; Dahl and Johnson, 1991; Frayer et al., 

1983). 

References 

Cowardin, L.M., V. Carter, F.C. Golet, and E.T. LaRoe. 1979. Classification of wetlands and deepwater 

habitats of the United States. FWS/OBS–79/31. Washington, DC: U.S. Department of the Interior, Fish 

and Wildlife Service. <http://library.fws.gov/FWS-OBS/79_31.pdf> 

Dahl, T.E. 2006. Status and trends of wetlands in the conterminous United States 1998 to 2004. 

Washington, DC: U.S. Department of the Interior, Fish and Wildlife Service. 

<http://wetlandsfws.er.usgs.gov/status_trends/> 

Dahl, T.E. 2000. Status and trends of wetlands in the conterminous United States 1986 to 1997. 

Washington, DC: U.S. Department of the Interior, U.S. Fish and Wildlife Service. 

<http://wetlandsfws.er.usgs.gov/status_trends/> 

Dahl, T.E. 1990. Wetlands losses in the United States 1780s to 1980s. Washington, DC: U.S. Department 

of the Interior, Fish and Wildlife Service. <http://wetlandsfws.er.usgs.gov/status_trends/> 

Dahl, T.E., and C.E. Johnson. 1991. Status and trends of wetlands in the conterminous United States, 

mid-1970s to mid-1980s. Washington, DC: U.S. Department of the Interior, U.S. Fish and Wildlife 

Service. <http://wetlandsfws.er.usgs.gov/status_trends/> 

Frayer, W.E., T.J. Monahan, D.C. Bowden, and F.A. Graybill. 1983. Status and trends of wetlands and 

deepwater habitats in the conterminous United States, 1950s to 1970s. Ft. Collins, CO: Colorado State 

University. <http://wetlandsfws.er.usgs.gov/status_trends/> 

Mitsch, W.J., and J.G. Gosselink. 1986. Wetlands. New York, NY: Van Nostrand Reinhold Company Inc. 

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3.4.3 Discussion 

What This Indicator Says About Trends in the Extent and Condition of 

Wetlands and Their Effects on Human Health and the Environment 

Wetland extent in the contiguous 48 states is substantially lower than it was prior to widespread European 

settlement and it generally continued to decline over the last 50 years (Wetlands indicator, p. 3-53). The 

rate of loss of wetlands overall and for most types of wetlands has slowed over time, however, and since 

1998 the overall extent of wetlands has actually increased. Not all types of wetlands have experienced the 

same rate of losses or overall percent losses. For example, freshwater shrub wetlands actually increased 

over the last 50 years—providing evidence of wetland conversion, most likely from forested wetlands to 

shrub. The nation has also seen a steady increase in acreage of freshwater ponds, which account for a 

substantial portion of the recent gains in overall wetland acreage. 

This indicator also confirms the role of many of the stressors described in Section 3.4.1. Over the last 

decade, development, forestry, and conversion to deepwater (e.g., marsh to open water) have led to losses 

in wetland extent, while agricultural areas have experienced overall gains in wetland acreage. The other 

source of new wetland acreage is from the “other” land use category, which reflects the growing 

importance of constructed and restored wetlands, including ponds associated with golf courses and 

residential development. 

While this indicator does not directly quantify the condition of the nation’s wetlands, it suggests that the 

condition of many wetlands may be impacted. As discussed in Section 3.4.1, extent is only a partial 

surrogate for condition because wetland loss can increase the stress on those wetlands that remain, while 

decreasing their connectivity. Thus, the overall decline in extent over the last 50 years suggests the 

potential for substantial ecological impacts such as habitat loss and increased flood impacts. Changes in 

the extent of different types of wetlands also suggest changes in condition. Shallow ponds, which 

constitute a large fraction of the recent gains in wetland acreage, will not perform the same range and type 

of environmental functions as the vegetated wetlands that disappeared between the 1950s and the 1990s. 

Similarly, evidence of wetland conversion indicates that even if extent is no longer declining rapidly, 

changes in wetland structure and function are still occurring. In the past, studies have shown that wetlands 

that have been created to mitigate for wetland losses have not yet provided the same functions and values 

of the wetlands that were lost.18, 19 

Limitations, Gaps, and Challenges 

By relying on aerial imagery and statistical surveying techniques, the Wetlands indicator (p. 3-53) 

provides a national estimate without an impractical number of samples. However, a limitation to this 

survey is that it may omit or undercount certain types of wetlands, including forested wetlands—which 

are difficult to photo-interpret—and ephemeral or well-drained agricultural wetlands, which are not 

necessarily obvious to the surveyor but are particularly threatened by development. This indicator also 

18 National Research Council. 2001. Compensating for wetland losses under the Clean Water Act. Washington, DC: 

National Academy Press. <http://www.nap.edu/books/0309074320/html/> 

19 Mack, J.J., and M. Micacchion. 2006. An ecological assessment of Ohio mitigation banks: vegetation, 

amphibians, hydrology, and soils. Ohio EPA Technical Report WET/2006-1. Columbus, OH: Ohio Environmental 

Protection Agency. <http://www.epa.state.oh.us/dsw/wetlands/WetlandBankReport.html> 

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does not include wetland parcels less than about 1 acre, which become more critical as larger wetlands are 

fragmented into smaller pieces. 

Wetland condition poses a larger challenge for assessment. While the Wetlands indicator (p. 3-53) 

provides information that can be used to infer potential wetland condition, it does not explicitly measure 

condition—in part because condition is difficult to quantify. Condition is made up of many different 

attributes, and each wetland has its own unique baseline condition, with a unique hydrologic setting and 

combination of plant and animal species. Some studies have quantified regional changes in specific 

stressors; however, national indicators would have to bring together many regional datasets and cover 

many different aspects of condition in order to be truly comprehensive. The lack of such national-scale 

information is currently a gap in addressing the question of wetland condition. Potential human health 

effects associated with wetland extent and condition are also difficult to quantify, and there are no ROE 

indicators on this topic. 

Another information gap concerns the spatial patterns of wetland change, which are not documented in 

the existing national data. Are most large wetlands being left intact? Are human activities threatening to 

fragment larger wetlands into smaller pieces that are less connected and more isolated, and therefore less 

able to perform the desired ecological functions? Data on patterns of wetland loss—e.g., fragmentation 

and edge effects—would be a useful complement to the existing data on overall losses and gains. 

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3.5 WHAT ARE THE TRENDS IN EXTENT AND CONDITION OF COASTAL 

WATERS AND THEIR EFFECTS ON HUMAN HEALTH AND THE 

ENVIRONMENT? 

3.5.1 Introduction 

Coastal waters are one of the nation’s most important natural resources, valued for their ecological 

richness as well as for the many human activities they support. As the interface between terrestrial 

environments and the open ocean, coastal waters encompass many unique habitats, such as estuaries, 

coastal wetlands, seagrass meadows, coral reefs, mangrove and kelp forests, and upwelling areas.20 

Coastal waters support many fish species for at least part of their life cycle, offering some of the most 

productive fisheries habitats in the world. These waters also provide breeding habitat for 85 percent of 

U.S. waterfowl and other migratory birds (largely in coastal wetlands),21 and support many other 

organisms with high public visibility (e.g., marine mammals, corals, and sea turtles) or unique ecological 

significance (e.g., submerged aquatic vegetation). For humans, coastal waters provide opportunities for 

tourism and recreation, and they contribute to the economy through transportation, fisheries, and mining 

and utilities.22 Lands adjacent to the coast are highly desirable places for people to live, and represent the 

most densely developed areas in the nation.23 

Extent and condition are two key variables in assessing coastal waters and their ability to serve ecological 

and human needs. The extent of coastal waters—i.e., the spatial area—is particularly important in terms 

of the extent of specific types of coastal waters, such as coastal wetlands. The condition of coastal waters 

reflects a group of interrelated physical, chemical, biological, and ecological attributes. For example, 

nutrient levels should be sufficient to support the food web but not so high as to cause eutrophication, 

while toxic chemical contaminants in water and sediment may pose a threat to aquatic organisms or 

accumulate in the food web. Of particular concern to human health are contaminants in consumable fish 

and shellfish—a topic discussed separately in Section 3.8. Other key aspects of condition include levels of 

pathogens and organisms that produce biotoxins—which may pose a risk to human health through aquatic 

recreation or contaminated fish and shellfish, and which may impact the environment by injuring native 

populations. Also important is the degree to which native plant and animal populations are healthy and 

their habitats intact. 

Many factors can affect the extent of coastal waters. For example, the extent of coastal wetlands may be 

influenced by natural events such as erosion or storms, or by human activities such as draining or filling 

wetlands for development. Natural processes can change the shape of a coastline, with wave action 

20 U.S. EPA. 2004. National coastal condition report II. EPA/620/R-03/002. Washington, DC. 

<http://www.epa.gov/owow/oceans/nccr/2005/index.html> 

21 U.S. EPA. 2004. National coastal condition report II. EPA/620/R-03/002. Washington, DC. 

<http://www.epa.gov/owow/oceans/nccr/2005/index.html> 

22 National Oceanic and Atmospheric Administration. 2005. Economic statistics for NOAA. May 2005—fourth 

edition. U.S. Department of Commerce. <http://www.publicaffairs.noaa.gov/pdf/economic-statistics2005.pdf> 

23 National Oceanic and Atmospheric Administration. 2004. Population trends along the coastal United States: 

1980–2008. Coastal trends report series. Silver Spring, MD: U.S. Department of Commerce, National Ocean 

Service. 

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eroding some areas while building up sediment in others, and rivers depositing sediments at their mouth. 

Human stressors can alter these patterns—for example, through the construction of seawalls or barriers or 

through the channeling of rivers, which can lead to subsidence in coastal areas that would otherwise be 

naturally replenished by sediments. 

Changes in extent may in turn affect the condition of coastal waters. For example, beach erosion and 

coastal wetland loss can also affect contaminant and sediment levels, nutrient cycling, and the condition 

of spawning and feeding grounds for fish, shellfish, and other coastal species. As described in Section 

3.4.1, the loss of some wetlands can also affect the condition of the wetlands that remain. 

Other stressors to the condition of coastal waters include nutrients, pathogens, and chemical 

contaminants, which may pose risks to ecological systems or to human health. Nutrients and pathogens 

occur naturally, but their abundance can be increased by human activities along the coast or in upstream 

watersheds that ultimately discharge to coastal waters. Major sources include urban and suburban storm 

water, agricultural runoff, and sewage discharge or overflows. Chemical contaminants may come from 

these same sources, as well as from industrial activities that discharge treated wastewaters and from 

atmospheric deposition of airborne pollutants. 

Several other stressors can affect the quality of habitat and the status of native plant and animal 

populations. For example, many species are sensitive to temperature and salinity, which can be influenced 

by changes in weather patterns or the condition of freshwater inputs. Salinity is particularly important in 

estuaries, where species may depend on a steady, reliable flow of fresh water. Another factor affecting the 

status of native communities is the presence and abundance of non-indigenous species—particularly 

invasive species that can kill or crowd out native populations, or otherwise alter coastal watersheds. 

Populations of fish, shellfish, marine mammals, and other species used by humans also may be affected 

by overharvesting. 

In many cases, stressors that affect coastal condition are interrelated. For example, excess nutrients can 

cause algal blooms (and subsequent decay) that result in low dissolved oxygen (DO) and reduced water 

clarity—the chain of events known as eutrophication. Temperature and salinity can also influence algal 

blooms. Some algae, such as “red tide,” produce toxins that pose risks to humans. 

3.5.2 ROE Indicators 

Five National Indicators and three Regional Indicators characterize the extent and condition of coastal 

waters. National Indicators describe sediment quality, benthic community condition, contamination in fish 

tissue, and several aspects of coastal water quality, as well as trends in the extent of marine and estuarine 

wetlands. The Regional Indicators characterize trends in harmful algal blooms, the extent of areas with 

low dissolved oxygen (i.e., hypoxia) and the extent of submerged aquatic vegetation (SAV). These 

Regional Indicators reflect conditions in three important and unique coastal water bodies: the Gulf of 

Mexico, Long Island Sound, and the Chesapeake Bay. 

The National Indicator on wetland extent is based on data gathered from aerial and ground surveys 

conducted as part of the U.S. Fish and Wildlife Service’s Wetlands Status and Trends study, a long-term 

statistical sampling effort. The other four National Indicators are derived from EPA’s second National 

Coastal Condition Report (NCCR II), which involved probabilistic surveys designed to represent 100 

percent of estuarine acreage in the contiguous 48 states and Puerto Rico. In addition to national totals, 

these four indicators also present data by EPA Region. The Regional Indicator on trends in hypoxia 

reflects data from two long-term water sampling programs, while the indicator on SAV is based on aerial 

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imagery. The Harmful Algal Blooms indicator reflects water sampling guided by satellite imagery that 

can detect blooms. 

Table 3.5.1. ROE Indicators of the Trends in Extent and Condition of Coastal Waters and their 

Effects on Human Health and the Environment 

NATIONAL INDICATORS LOCATION 

Wetland Extent, Change, and Sources of Change 3.4.2 – p. 3-53 

Trophic State of Coastal Waters (N/R) 3.5.2 – p. 3-62 

Coastal Sediment Quality (N/R) 3.5.2 – p. 3-67 

Coastal Benthic Communities (N/R) 3.5.2 – p. 3-71 

Coastal Fish Tissue Contaminants (N/R) 3.8.2 – p. 3-103 

REGIONAL INDICATORS LOCATION 

Submerged Aquatic Vegetation in the Chesapeake Bay 3.5.2 – p. 3-74 

Hypoxia in the Gulf of Mexico and Long Island Sound 3.5.2 – p. 3-77 

Harmful Algal Bloom Outbreaks Along the Western Florida Coastline 3.5.2 – p. 3-81 

N/R = National Indicator displayed at EPA Regional scale 

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INDICATOR: Trophic State of Coastal Waters 

While many water pollutants can lead to decreases in coastal water quality, four interlinked components 

related to trophic state are especially critical: nutrients (nitrogen and phosphorus), chlorophyll-a,

dissolved oxygen (DO), and water clarity. Trophic state generally refers to aspects of aquatic systems 

associated with the growth of algae, decreasing water transparency, and low oxygen levels in the lower 

water column that can harm fish and other aquatic life. Nitrogen is usually the most important limiting 

nutrient in estuaries, driving large increases of microscopic phytoplankton called “algal blooms,” but 

phosphorus can become limiting in coastal systems if nitrogen is abundant in a bioavailable form (U.S. 

EPA, 2003). Nitrogen and phosphorus can come from point sources, such as wastewater treatment plants 

and industrial effluents, and nonpoint sources, such as runoff from farms, over-fertilized lawns, leaking 

septic systems, and atmospheric deposition. Chlorophyll-a is a surrogate measure of algal abundance. 

Chlorophyll-a levels are increased by nutrients and decreased by filtering organisms (e.g., clams, mussels, 

or oysters). High concentrations of chlorophyll-a indicate overproduction of algae, which can lead to algal 

scums, fish kills, and noxious odors (U.S. EPA, 2004). Low dissolved oxygen levels and decreased clarity 

caused by algal blooms or the decay of organic matter from the watershed are stressful to estuarine 

organisms. Reduced water clarity (usually measured as the amount and type of light penetrating water to a 

depth of one meter) also can be caused by storm-related events that cause erosion or mixing from the 

sediments, and can impair the normal growth of algae and other submerged aquatic vegetation. 

This indicator, developed as part of EPA’s Coastal Condition Report, is based on an index constructed 

from probabilistic survey data on five components: dissolved inorganic nitrogen, dissolved inorganic 

phosphorus, chlorophyll-a, dissolved oxygen in bottom or near-bottom waters (where benthic life is most 

likely to be affected), and water clarity (U.S. EPA, 2004). The survey, part of EPA’s National Coastal 

Assessment (NCA), was designed to provide a national picture of water quality by sampling sites in 

estuarine waters throughout the contiguous 48 states and Puerto Rico. Each site was sampled once during 

the period 1997-2000, within an index period from July to September. The indicator reflects average 

condition during this index period. 

Reference conditions were established for each EPA Region for nutrients, water clarity, and chlorophyll-a

because key factors like sediment load, mixing parameters, and ecosystem sensitivity naturally vary from 

one Region to the next. A single national reference range of 2-5 mg/L was used for dissolved oxygen, 

because concentrations below 2 mg/L are almost always harmful to many forms of aquatic life, and 

concentrations above 5 mg/L seldom are (Diaz and Rosenberg, 1995; U.S. EPA, 2000). The process of 

classifying individual sites varies by Region and is described in detail, along with the regional reference 

conditions, in U.S. EPA, 2004 (pp. 19-20). 

The overall water quality index is a compilation of the five components. For each site, the index is rated 

high if none of the five components of the index received a score that would be considered 

environmentally unfavorable (high nitrogen, phosphorus, or chlorophyll-a levels or low DO or water

clarity), and no more than one component was rated moderate. Overall water quality is low if more than 

two components received the most unfavorable rating. All other sites receive a moderate index score. If 

two or more components are missing, and the available components do not suggest a moderate or low 

index rating, the site is classified as “unsampled.” Data from the individual sites were expanded from the 

probability sample to provide unbiased estimates of the water quality index and each of its components 

for each EPA Region and for the entire nation. 
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What the Data Show 

According to the index, 40 percent of estuarine surface area nationwide exhibited high water quality over 

the period 1997-2000, 11 percent had low water quality, and the remaining 49 percent was rated moderate 

(Exhibit 3-20). Scores vary considerably among EPA Regions, ranging from high water quality in 71 

percent of estuarine area in Region 1 to less than 10 percent in Regions 2 and 3. Only one EPA Region 

had low water quality in more than 15 percent of its estuarine area (EPA Region 3, with 36 percent). 

These percentages do not include the Great Lakes or the hypoxic zone in offshore Gulf Coast waters (see 

the Hypoxia in Gulf of Mexico and Long Island Sound indicator, p. 3-77). 

Nitrogen concentrations were low in 82 percent of estuarine area and high in 5 percent nationwide, and 

were low in a majority of the estuarine area in all but one EPA Region (Exhibit 3-21). Regions 2 and 3 

had the largest percentage of area with high concentrations (15 percent and 16 percent, respectively); 

several other EPA Regions had no areas with high concentrations. 

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Phosphorus concentrations were low in 53 percent of estuarine area and high in 9 percent nationwide 

(Exhibit 3-22). Region 9 had the largest proportion of area exceeding reference conditions (52 percent), 

while Region 10 had the least (none). 

Chlorophyll-a concentrations were low in 51 percent and high in 8 percent of estuarine area nationwide 

(Exhibit 3-23). Region 3 had the largest percentage of area exceeding reference conditions (27 percent); 

all other EPA Regions had 10 percent or less in this category. 

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Bottom-water dissolved oxygen was above 5 mg/L in over three-fourths of the nation’s estuarine area and 

below 2 mg/L in only 4 percent (Exhibit 3-24). While effects vary with temperature and salinity, as a 

general rule, concentrations of DO above 5 mg/L are considered supportive of marine life, concentrations 

below 5 mg/L are potentially harmful, and concentrations below 2 mg/L—a common threshold for 

hypoxia—are associated with a wider range of harmful effects (e.g., some juvenile fish and crustaceans 

that cannot leave the area may die). Region 3 had the greatest proportion of estuarine area with low DO 

(21 percent), while four EPA Regions had no area below 2 mg/L. 

Water clarity exceeded reference conditions (i.e., higher clarity) in 62 percent of the nation’s estuarine 

area, while low water clarity was observed in 25 percent of estuarine area (Exhibit 3-25). Region 3 had 

the largest proportion of area with low clarity (43 percent), while Region 1 had the least (none). 

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Indicator Limitations 

• The indicator does not include data from the Great Lakes, which are monitored using a 

different index design. The coastal areas of Hawaii and a portion of Alaska have been 

sampled, but the data had not yet been assessed at the time this indicator was compiled. Data 

are also not available for the U.S. Virgin Islands and the Pacific territories. 

• Trend data are not yet available for this indicator. Because of differences in methodology, the 

data presented here are not comparable with data that appeared in EPA’s first National 

Coastal Condition Report (NCCR I). The data presented here will serve as a baseline for 

future surveys. 

• The National Coastal Assessment surveys measure dissolved oxygen conditions only in 

estuarine waters and do not include observations of dissolved oxygen concentrations in 

offshore coastal shelf waters, such as the hypoxic zone in Gulf of Mexico shelf waters. 

• At each sample location, the components of this indicator may have a high level of temporal 

variability. This survey is intended to characterize the typical distribution of water quality 

conditions in coastal waters during an index period from July through September. It does not 

consistently identify the “worst-case” condition for sites experiencing occasional or 

infrequent hypoxia, nutrient enrichment, or decreased water clarity at other times of the year. 

Data Sources 

This indicator is based on an analysis published in EPA’s National Coastal Condition Report II (U.S. 

EPA, 2004). Summary data by EPA Region have not been published, but were provided by EPA’s 

National Coastal Assessment program. Underlying sampling data are housed in EPA’s National Coastal 

Assessment database (U.S. EPA, 2005) (http://www.epa.gov/emap/nca/html/data/index.html). 

References 

Diaz, R.J., and R. Rosenberg. 1995. Marine benthic hypoxia: a review of its ecological effects and the 

behavioral responses of benthic macrofauna. Oceanogr. Mar. Biol. Ann. Rev. 33:245-303. 

U.S. EPA. 2005. EMAP national coastal database. Accessed 2005. 

<http://www.epa.gov/emap/nca/html/data/index.html> 

U.S. EPA. 2004. National coastal condition report II. EPA/620/R-03/002. Washington, DC. 

<http://www.epa.gov/owow/oceans/nccr/2005/index.html> 

U.S. EPA. 2003. Mid-Atlantic integrated assessment, MAIA—estuaries 1997-98, summary report. 

EPA/620/R-02/003. Narragansett, RI. 

U.S. EPA. 2000. Ambient aquatic life water quality criteria for dissolved oxygen (saltwater): Cape Cod to 

Cape Hatteras. EPA/822/R-00/12. 

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INDICATOR: Coastal Sediment Quality 

Contaminated sediments can pose an immediate threat to benthic organisms and an eventual threat to 

entire estuarine ecosystems. Sediments can be resuspended by anthropogenic activities, storms or other 

natural events; as a result, organisms in the water column can be exposed to contaminants, which may 

accumulate through the food web and eventually pose health risks to humans (U.S. EPA, 2004a). 

There are several ways to measure sediment quality. Sediments can be assessed in terms of their toxicity 

to specific organisms in bioassays, or in terms of the levels of contaminants that are present. Sediment 

quality also can be inferred by assessing the condition of benthic communities, which largely reflect the 

quality of the sediments in which they live (although other stressors may be reflected as well). To 

generate a more complete picture of sediment quality, scientists frequently use several of these measures 

together.

This indicator presents data on sediment toxicity and contaminant levels. The data are from probabilistic 

surveys conducted as part of EPA’s National Coastal Assessment (NCA) and presented in EPA’s second 

National Coastal Condition Report (U.S. EPA, 2004b). The survey was designed to provide a national 

picture of sediment quality by sampling sites in estuarine waters throughout the contiguous 48 states and 

Puerto Rico. Each site was sampled once during the period 1997-2000, within an index period from July 

to September. The indicator reflects average condition during this index period. 

Sediment toxicity is typically determined using bioassays that expose test organisms to sediments and 

evaluate their effects on the organisms’ survival. For this indicator, toxicity was determined using a 10-

day static test on the benthic amphipod Ampelisca abdita, which is commonly used as a screening tool to 

identify sediments that pose sufficient concern to warrant further study. Sediments were classified as 

“potentially toxic” if the bioassays resulted in greater than 20 percent mortality (a reference condition), or 

“not likely toxic” if the bioassays resulted in 20 percent mortality or less (U.S. EPA, 2004c). 

Contaminant concentrations do not directly reflect toxicity because toxicity also depends on 

contaminants’ bioavailability, which is controlled by pH, particle size and type, organic content, and other 

factors (e.g., mercury vs. methylmercury). Contaminant concentrations are a useful screening tool for 

toxicity, however, when compared with concentrations known to cause particular effects on benthic life. 

For this indicator, sediment samples were homogenized and analyzed for nearly 100 contaminants, 

including 25 PAHs, 22 PCBs, 25 pesticides, and 15 metals, using standard wet chemistry and mass 

spectroscopy. The observed concentrations were then compared with “effects range median” (ERM) 

values established through an extensive review of toxicity tests involving benthic organisms, mostly 

Ampelisca (Long et al., 1995). For each contaminant, the ERM represents the concentration at which 

there is a 50 percent likelihood of adverse effects to an organism, based on experimental data. For this 

indicator, a site was rated “potentially toxic” if one or more contaminants exceeded an ERM value. In 

practice, about 25 percent of samples that exceed one ERM also cause more than 20 percent mortality in 

the Ampelisca bioassay (Long, 2000). 

Benthic community condition also can be a useful indication of sediment quality, particularly in terms of 

chronic or community effects that would not be captured in an acute exposure bioassay. The NCA 

evaluated estuarine sites for several aspects of benthic community condition, and these results are 

presented as a separate ROE indicator (Coastal Benthic Communities, p. 3-71). 
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What the Data Show 

Nationwide, 6 percent of coastal sediments were rated “potentially toxic” based on the Ampelisca toxicity 

screening assay, although there was considerable variability from one EPA Region to the next (Exhibit 3-

26). In Region 9, nearly 100 percent of estuarine area exhibited low sediment toxicity, while in some 

other EPA Regions, as much as 20 percent of estuarine sediments were “potentially toxic.” Data for 

Region 6 are inconclusive because more than half of the Region’s estuarine area was not sampled. 

Nationally, contaminants were present at “potentially toxic” levels in 7 percent of estuarine sediments for 

which contamination data were available (Exhibit 3-27). There was considerable variability in sediment 

contamination from one EPA Region to the next, with Region 4 showing the largest proportion of 

estuarine area with sediments not likely to be toxic (99.9 percent), and Region 2 showing the largest 

proportion with “potentially toxic” sediments (24.4 percent). 

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Although the two figures suggest that a similar percentage of the nation’s estuarine sediments are 

“potentially toxic,” the original data source reports very little correlation between sites that caused >20 

percent mortality in the Ampelisca bioassay and sites where one or more contaminants exceeded the ERM 

(U.S. EPA, 2004b). It is not unusual to find a lack of correlation—particularly in cases where sediment 

contaminants are neither highly concentrated nor completely absent—in part because some toxic 

chemicals may not be bioavailable, some may not be lethal, and not all potentially toxic chemicals are 

analyzed (see O’Connor et al., 1998, and O’Connor and Paul, 2000). These results underscore the utility 

of a combined approach to screen for potentially toxic sediments. 

Indicator Limitations 

• The coastal areas of Hawaii and a portion of Alaska have been sampled, but the data had not 

yet been assessed at the time this indicator was compiled. Data are also not available for the 

U.S. Virgin Islands and the Pacific territories. 

• Trend data are not yet available for this indicator. Because of differences in methodology, the 

data presented here are not comparable with data that appeared in EPA’s first National 

Coastal Condition Report (NCCR I). The data presented here will serve as a baseline for 

future surveys. 

• Sample collection is limited to an index period from July to September. It is not likely that 

contaminant levels vary from season to season, however. 

• The Ampelisca bioassay is a single-organism screening tool, and the ERMs are general 

screening guidelines based largely on toxicity data from Ampelisca. Thus, these measures do 

not necessarily reflect the extent to which sediments may be toxic to the full range of biota 

(including microbes and plants) that inhabit a particular sampling location. 

• The Ampelisca bioassay tests only for short-term, not long-term, exposure. Both screening 

tests characterize sediments in terms of their effects on benthic organism mortality. This 

indicator does not capture other effects of sediment contaminants on benthic organisms, such 

as disease, stress, and reproductive effects. 

• This indicator cannot be compared quantitatively with indicators that use other types of 

contaminant guidelines. For example, the Pesticides in Agricultural Streams indicator (p. 3-

32) uses thresholds intended to be protective of aquatic life with a margin of safety, instead of 

thresholds shown to cause biological effects (e.g., ERMs). The ERM approach also is not 

directly comparable with other sediment contaminant approaches, such as EPA’s equilibrium 

partitioning (EqP) benchmarks. 

Data Sources 

This indicator is based on an analysis published in EPA’s National Coastal Condition Report II (U.S. 

EPA, 2004). Summary data by EPA Region have not been published, but were provided by EPA’s 

National Coastal Assessment program. Underlying sampling data are housed in EPA’s National Coastal 

Assessment database (U.S. EPA, 2005) (http://www.epa.gov/emap/nca/html/data/index.html). 

References 

Long, E.R. 2000. Degraded sediment quality in U.S. estuaries: a review of magnitude and ecological 

applications. Ecol. Appl. 10(2):338-349. 

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Long, E.R., D.D. MacDonald, L. Smith, and F.D. Calder. 1995. Incidence of adverse biological effects 

within ranges of chemical concentrations in marine and estuarine sediments. Environ. Manage. 19:81-97. 

O’Connor, T.P., K.D. Daskalakis, J.L. Hyland, J.F. Paul, and J.K. Summers. 1998. Comparisons of 

sediment toxicity with predictions based on chemical guidelines. Environ. Toxicol. Chem. 17(3):468-471. 

O’Connor, T.P., and J.F. Paul. 2000. Misfit between sediment toxicity and chemistry. Mar. Pollut. Bull. 

40(1):59-64. 

U.S. EPA. 2005. EMAP national coastal database. Accessed 2005. 

<http://www.epa.gov/emap/nca/html/data/index.html> 

U.S. EPA. 2004a. Contaminated sediment in water. <http://www.epa.gov/waterscience/cs/aboutcs> 

U.S. EPA. 2004b. National coastal condition report II. EPA/620/R-03/002. Washington, DC. 

<http://www.epa.gov/owow/oceans/nccr/2005/index.html> 

U.S. EPA. 2004c. The incidence and severity of sediment contamination in surface waters of the United 

States—national sediment quality survey. Second edition. Washington, DC. 

<http://epa.gov/waterscience/cs/report/2004/nsqs2ed-complete.pdf> 

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Benthic communities are largely composed of macroinvertebrates, such as annelids, mollusks, and 

crustaceans. These organisms inhabit the bottom substrates of estuaries and play a vital role in 

maintaining sediment and water quality. They also are an important food source for bottom-feeding fish, 

invertebrates, and birds. Communities of benthic organisms are important indicators of environmental 

stress because they are particularly sensitive to pollutant exposure (Holland et al., 1987). This sensitivity 

arises from the close relationship between benthic organisms and sediments—which can accumulate 

environmental contaminants over time—and the fact that these organisms are relatively immobile, which 

means they receive prolonged exposure to any contaminants in their immediate habitat (Sanders et al., 

1980; Nixon et al., 1986). 

This indicator is based on a multi-metric benthic communities index that reflects overall species diversity 

in estuarine areas throughout the contiguous United States (adjusted for salinity, if necessary) and, for 

some regions, the presence of pollution-tolerant and pollution-sensitive species (e.g., Weisberg et al., 

1997; Engle and Summers, 1999; U.S. EPA, 2004). 

The benthic community condition at each sample site 

is given a high score if the index exceeds a particular

threshold (e.g., has high diversity or populations of 

many pollution-sensitive species), a low score if it falls 

below the threshold conditions, and a moderate score if 

it falls within the threshold range. The exact structure

of the index and the threshold values vary from one 

region to another, but comparisons between predicted 

and observed scores based on expert judgment are used 

to ensure that the classifications of sites from one 

region to another are consistent (see U.S. EPA, 2004, 

p. 15). Data were collected using probability samples, 

so the results from the sampling sites provide unbiased 

estimates of the distribution of index scores in 

estuaries throughout each region. 

The data for this indicator are from probabilistic 

surveys conducted as part of EPA’s National Coastal 

Assessment (NCA) and presented in EPA’s second 

National Coastal Condition Report (U.S. EPA, 2004b).

The survey was designed to provide a national picture

of coastal benthic community condition by sampling 

sites in estuarine waters throughout the contiguous 48 

states and Puerto Rico. Each site was sampled once 

during the period 1997-2000, within an index period 

from July to September. The indicator reflects average 

condition during this index period. 

What the Data Show

Nationally, 70 percent of the sampled estuarine area 

had a high benthic communities index score, with 13 

percent in the moderate range and 17 percent scoring 
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low (Exhibit 3-28). Condition varied somewhat by EPA Region, with high index scores ranging from 51 

percent of the estuarine area in Region 6 to 91 percent in Region 10. Region 3 had the largest proportion 

of estuarine area rated low (27 percent), while Region 10 had the lowest (4 percent). In the figure, the 

portion of the estuarine area not represented by the sample is noted for each Region. 

The National Coastal Condition Report found that many of the sites with low benthic community 

condition also showed impaired water quality or sediment condition—which is not surprising given the 

extent to which these stressors and effects are related. Of the 17 percent of national estuarine area rated 

low on the benthic communities index, 38 percent also exhibited degraded sediment quality, 9 percent 

exhibited degraded water quality (U.S. EPA, 2004), and 33 percent exhibited degraded sediment and 

water quality. 

Indicator Limitations 

• The indicator does not include data from the Great Lakes, which are monitored using a 

different index design. The coastal areas of Hawaii and a portion of Alaska have been 

sampled, but the data had not yet been assessed at the time this indicator was compiled. 

• Trend data are not yet available for this indicator. Because of differences in methodology, the 

data presented here are not comparable with data that appeared in EPA’s first National 

Coastal Condition Report (NCCR I). The data presented here will serve as a baseline for 

future surveys. 

• Benthic indices for the Northeast, West, and Puerto Rico do not yet include measures of 

pollution-tolerant or pollution-sensitive species. Although species diversity has the largest 

impact on index scores in the other regions, index values could change in the future as these 

components are added to the index values for these regions. 

• Sample collection is limited to an index period from July to September. Further, because 

benthic communities can be strongly influenced by episodic events, trawling, or climate 

perturbations, this indicator may not reflect the full range of conditions that occur at each 

sampling location throughout these months. 

Data Sources 

This indicator is based on an analysis published in EPA’s National Coastal Condition Report II (U.S. 

EPA, 2004). Summary data by EPA Region have not been published, but were provided by EPA’s 

National Coastal Assessment program. Underlying sampling data are housed in EPA’s National Coastal 

Assessment database (U.S. EPA, 2005) (http://www.epa.gov/emap/nca/html/data/index.html). 

References 

Engle, V.D., and J.K. Summers. 1999. Refinement, validation, and application of a benthic condition 

index for northern Gulf of Mexico estuaries. Estuaries 22(3A):624-635. 

Holland, A.F., A. Shaughnessy, and M.H. Heigel. 1987. Long-term variation in mesohaline Chesapeake 

Bay benthos: spatial and temporal patterns. Estuaries 10:227-245. 

Nixon, S.W., C.D. Hunt, and B.L. Nowicki. 1986. The retention of nutrients (C, N, P), heavy metals (Mn, 

Cd, Pb, Cu), and petroleum hydrocarbons by Narragansett Bay. In: Lasserre, P., and J.M. Martin, eds. 

Biogeochemical processes at the land-sea boundary. New York, NY: Elsevier. pp. 99-122. 

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Sanders, H.L., J.F. Grassle, G.R. Hampson, L.S. Morse, S. Gerner-Price, and C.C. Jones. 1980. Anatomy 

of an oil spill: long-term effects from the grounding of the barge Florida off West Falmouth, 

Massachusetts. J. Mar. Res. 38:265-380. 

U.S. EPA. 2005. EMAP national coastal database. Accessed 2005. 

<http://www.epa.gov/emap/nca/html/data/index.html> 

U.S. EPA. 2004. National coastal condition report II. EPA/620/R-03/002. Washington, DC. 

<http://www.epa.gov/owow/oceans/nccr/2005/index.html> 

Weisberg, S.B., J.A. Ranasinghe, D.D. Dauer, L.C. Schnaffer, R.J. Diaz, and J.B. Frithsen. 1997. An 

estuarine benthic index of biotic integrity (B-IBI) for Chesapeake Bay. Estuaries 20(1):149-158 

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1 INDICATOR: Submerged Aquatic Vegetation in the Chesapeake Bay 

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Submerged Aquatic Vegetation (SAV) is important to the Chesapeake Bay aquatic ecosystem. SAV 

supports the Bay’s health by: 

• generating food and habitat for waterfowl, fish, shellfish and invertebrates; 

• adding oxygen to the water column during photosynthesis; 

• filtering and trapping sediment that otherwise would bury benthic organisms and cloud the 

water column; 

• inhibiting wave action that erodes shorelines; and 

• absorbing nutrients, such as nitrogen and phosphorus, that otherwise could fuel the growth of 

unwanted planktonic algae. 

The loss of SAV from shallow waters of the Chesapeake Bay, which was first noted in the early 1960s, is 

a widespread, well-documented problem (Batiuk et al., 2000). Review of aerial photographs taken from a 

number of sites taken between the mid-1930s and the mid-1960s suggests that current SAV acreage is less 

than half of that during the earlier period (Moore et al., 2004). 

Trends in the distribution and abundance of SAV over time are useful in understanding trends in water 

quality (Moore et al., 2004). Although other factors, such as climatic events and herbicide toxicity, may 

have contributed to the decline of SAV in the Bay, the primary causes are eutrophication and associated 

reductions in light availability (Batiuk et al., 2000). Like all plants, SAV needs sunlight to grow and 

survive. Two key stressors that impact the growth of SAV are suspended sediments and excess nutrient 

pollution. Suspended sediments—loose particles of clay and silt that are suspended in the water—make 

the water dingy and block sunlight from reaching the plants. Similarly, excess nutrients in the water fuel 

the growth of planktonic algae, which also block sunlight. 

This indicator presents the distribution of SAV in the Chesapeake Bay and its tributaries from 1978 to 

2005, as mapped from black and white aerial photographs. The surveys follow fixed flight routes to 

comprehensively survey all shallow water areas of the Bay and its tidal tributaries. Non-tidal areas are 

omitted from the survey. SAV beds less than 1 square meter in area are not included due to the limits of 

the photography and interpretation. Annual monitoring began in 1978; however, no surveys were 

conducted from 1979 to 1983 or in 1988. In years when the entire area could not be surveyed due to flight 

restrictions or weather events, acreages in the non-surveyed areas were estimated based on prior years’ 

surveys. 

What the Data Show

The extent of SAV in the Chesapeake Bay increased from 41,000 acres in 1978 to 78,000 acres in 2005 

(Exhibit 3-29). The extent of SAV declined to a minimum of 38,000 acres in 1984 and the maximum 

extent of SAV during this period was 90,000 acres, in 2002. The notable decline in SAV distribution 

between 2002 and 2003 appears to be the result of substantial reductions in widgeongrass populations in 

the lower and mid-bay regions. In addition to the large declines in widgeongrass, major declines in 

freshwater SAV species occurred in the upper portion the Potomac River and Susquehanna region. While 

populations of SAV appeared to be present in these segments very early in the growing season, persistent 
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turbidity resulting from rain occurring throughout the spring and summer may have contributed to a very 

early decline, well before Hurricane Isabel affected the Chesapeake Bay (Orth et al., 2004). The extent of 

SAV gradually increased again through 2004 and 2005. 

Indicator Limitations 

• There were no surveys in the years 1979-

1983 or in 1988. 

• The indicator includes some estimated 

data for years with incomplete 

photographic coverage. Spatial gaps in 

1999 occurred due to the inability to 

reliably photograph SAV following 

hurricane disturbance. Spatial gaps in 

2001 occurred due to flight restrictions 

near Washington D.C. after the September 

11th terrorist attacks. Other gaps occurred 

in 2003 due to adverse weather in the 

spring, summer, and fall (Hurricane 

Isabel). Acreage in the non-surveyed areas 

was estimated based on prior years’ 

surveys. 

• Photointerpretation methods changed over 

the course of this study. However, data 

have been adjusted to account for any 

methodological inconsistencies. 

• Extent is just one of the variables that can 

be used to measure the condition of SAV 

communities. Information on vegetation 

health and density would also provide useful information. 

Data Sources 

Data were obtained from the Chesapeake Bay Program, which has published a version of this indicator 

(Chesapeake Bay Program, 2006) along with a link to download the annual summary data presented in 

Exhibit 3-29 (http://www.chesapeakebay.net/pubs/statustrends/88-data-2002.xls). These acreage statistics 

are based on annual SAV distribution maps, which are available from the Virginia Institute of Marine 

Science (VIMS, 2006) (http://www.vims.edu/bio/sav/savdata.html). 

References 

Batiuk, R., P. Bergstrom, M. Kemp, E. Koch, L. Murray, C. Stevenson, R. Bartleson, V. Carter, N. 

Rybicki, J. Landwehr, C. Gallegos, L. Karrh, M. Naylor, D. Wilcox, K. Moore, S. Ailstock, and M. 

Teichberg. 2000. Chesapeake Bay submerged aquatic vegetation water quality and habitat-based 

requirements and restoration targets: a second technical synthesis. CBP/TRS 245/00. EPA/903/R-00/014. 

Annapolis, MD: U.S. EPA, Chesapeake Bay Program. 

<http://www.chesapeakebay.net/pubs/sav/savreport.pdf>. 

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Chesapeake Bay Program. 2006. Bay trends & indicators: bay grass abundance (baywide). 

<http://www.chesapeakebay.net/status.cfm?sid=88> 

Moore, K.A., D.J. Wilcox, B. Anderson, T.A. Parham, and M.D. Naylor. 2004. Historical analysis of 

SAV in the Potomac River and analysis of bay-wide historic SAV to establish a new acreage goal. Report 

to EPA Chesapeake Bay Program. 

<http://www.vims.edu/bio/sav/Final_SAV_Historical_Report_2004.pdf> 

Orth, R.J., D.J. Wilcox, L.S. Nagey, A.L. Owens, J.R. Whiting, and A. Serio. 2004. 2003 distribution of 

submerged aquatic vegetation in Chesapeake Bay and coastal bays. Report to EPA Chesapeake Bay 

Program. <http://www.vims.edu/bio/sav/sav03/index.html> 

VIMS. 2006. SAV distribution data archive. Virginia Institute of Marine Science. Accessed 2006. 

<http://www.vims.edu/bio/sav/savdata.html> 

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1 INDICATOR: Hypoxia in the Gulf of Mexico and Long Island Sound 

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Nutrient pollution is one of the most pervasive problems facing U.S. coastal waters, with more than half 

of the nation’s estuaries experiencing one or more symptoms of eutrophication (U.S. Commission on 

Ocean Policy, 2004; NRC, 2000; Bricker et al., 1999). One symptom is low levels of dissolved oxygen 

(DO), or hypoxia. Hypoxia can occur naturally, particularly in areas where natural physical and chemical 

characteristics (e.g., salinity or mixing parameters) limit bottom-water DO. The occurrence of hypoxia in 

shallow coastal and estuarine areas appears to be increasing, however, and is most likely accelerated by 

human activities (Vitousek et al., 1997; Jickells, 1998). 

This indicator tracks trends in hypoxia in the Gulf of Mexico and Long Island Sound, which are prime 

examples of coastal and estuarine areas experiencing hypoxia. For consistency, this indicator focuses on 

occurrences of DO below 2 mg/L, but actual thresholds for “hypoxia” and associated effects can vary 

over time and space. Hypoxia often is defined as a concentration of DO below saturation, and because 

saturation levels vary with temperature and salinity, the concentration that defines hypoxia will vary 

seasonally and geographically. Effects of hypoxia on aquatic life also vary, as some organisms are more 

sensitive to low DO than others. As a general rule, however, concentrations of DO above 5 mg/L are 

considered supportive of marine life, while concentrations below this are potentially harmful. At about 3 

mg/L, bottom fishes may start to leave the area, and the growth of sensitive species such as crab larvae is 

reduced. At 2.5 mg/L, the larvae of less sensitive species of crustaceans may start to die, and the growth 

of crab species is more severely limited. Below 2 mg/L, some juvenile fish and crustaceans that cannot 

leave the area may die, and below 1 mg/L, fish totally avoid the area or begin to die in large numbers 

(Howell and Simpson, 1994; U.S. EPA, 2000). 

The Gulf of Mexico hypoxic zone on the Texas-Louisiana Shelf is the largest zone of coastal hypoxia in 

the Western Hemisphere (CAST, 1999). It exhibits seasonally low oxygen levels as a result of 

complicated interactions involving excess nutrients carried to the Gulf by the Mississippi and Atchafalaya 

Rivers; physical changes in the river basin, such as channeling, construction of dams and levees, and loss 

of natural wetlands and riparian vegetation; and the stratification in the waters of the northern Gulf caused 

by the interaction of fresh river water and the salt water of the Gulf (CENR, 2000; Rabalais and Turner, 

2001). Increased nitrogen and phosphorus inputs from human activities throughout the basin support an 

overabundance of algae, which die and fall to the sea floor, depleting oxygen in the water as they 

decompose. Fresh water from the rivers entering the Gulf of Mexico forms a layer of fresh water above 

the saltier Gulf waters and prevents re-oxygenation of oxygen-depleted water along the bottom. 

In Long Island Sound, seasonally low levels of oxygen usually occur in bottom waters from mid-July 

though September, and are more severe in the western portions of the Sound, where the nitrogen load is 

higher and stratification is stronger, reducing mixing and re-oxygenation processes (Welsh et al., 1991). 

While nitrogen fuels the growth of microscopic plants that leads to low levels of oxygen in the Sound, 

temperature, wind, rainfall, and salinity can affect the intensity and duration of hypoxia. 

Data for the two water bodies are presented separately because they are collected through two different 

sampling programs, each with its own aims and technical approach. The Gulf of Mexico survey is 

designed to measure the extent of bottom-water hypoxia in the summer, with samples collected during a 

cruise that generally occurs over a five-day period in mid-to-late July (LUMCON, 2006). Samples are 

collected day and night along several transects designed to capture the overall extent of the hypoxic zone. 

The number of locations varies from 60 to 90 per year, depending on the length of the sampling cruise, 

the size of the hypoxic zone, logistical constraints, and the density of station locations. Long Island Sound 

sampling is designed to determine both the maximum extent and the duration of hypoxia (Connecticut 
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DEP, 2006a). Sampling is performed every month from October to May and every two weeks from June 

to September at a set of fixed locations throughout the Sound. All Long Island Sound samples are 

collected during the day. 

What the Data Show 

The size of the midsummer bottom-water hypoxia area 

(<2 mg/L DO) in the Northern Gulf of Mexico has 

varied considerably since 1985, ranging from 40 

square miles in 1988 (a drought year in the Mississippi 

Basin) to approximately 8,500 square miles in 2002 

(Exhibit 3-30). The unusually low areal extent in 2000 

also was associated with very low discharge from the 

Mississippi River (see the N and P Discharge from 

Large Rivers indicator, p. 3-28). In the latest year of 

sampling, 2005, the hypoxic zone measured over 

4,500 square miles, slightly smaller than the state of 

Connecticut (Exhibit 3-31). Over the full period of 

record (1985–2005), the area with DO <2.0 mg/L has 

averaged approximately 4,900 square miles. 

The maximum extent and duration of hypoxic events 

(<2 mg/L DO) in Long Island Sound also has varied 

considerably since 1987 (Exhibit 3-32). Since 1987, 

the largest area of DO less than 2 mg/L was 212 

square miles, which occurred in 1994; the smallest area, 2 square miles, occurred in 1997 (panel A). The 

shortest hypoxic event was 6 days in 1990 and the longest was 71 days, in 1989 (panel B). In 2005, the 

latest year for which data are available, the maximum area and duration of DO <2 mg/L in Long Island 

Sound were 95 square miles and 60 days, respectively, with the lowest DO levels occurring in the western 

end of the Sound (Exhibits 3-32 and 3-33). Between 1987 and 2005, the average annual maximum was 71 

square miles and 33 days. 

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Indicator Limitations 

 Gulf of Mexico 

• This indicator is based on a survey 

conducted over a five-day period when 

hypoxia is expected to be at its maximum 

extent. The indicator does not capture 

periods of hypoxia or anoxia (no oxygen 

at all) occurring at times other than the 

mid-summer surveys. 

• Because the extent of hypoxia is measured 

through a single mid-summer sampling 

cruise, duration cannot be estimated. 

• This indicator does not track vertical 

extent of hypoxia or anoxic volume. 

• Surveys usually end offshore from the 

Louisiana-Texas State line; in years when 

hypoxia extends onto the upper Texas 

coast, the spatial extent of hypoxia is 

underestimated. 

 Long Island Sound 

• Hypoxic or anoxic periods that may occur 

between the two-week surveys are not 

captured in the indicator. 

• Samples are taken in the 

daytime, approximately 

one meter off the bottom. 

This indicator does not 

capture oxygen conditions 

at night, which may be 

lower because of the lack 

of photosynthesis, or 

conditions near the 

sediment-water interface. 

Data Sources 

Gulf of Mexico data were provided by 

the Louisiana Universities Marine 

Consortium (LUMCON). Maps and 

summary data from the most recent 

Gulf of Mexico survey are published 

online (LUMCON, 2006). 

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Long Island Sound data were provided by the Connecticut Department of Environmental Protection’s 

Long Island Sound Water Quality Monitoring Program. Data on extent and duration of hypoxia have not 

been published, but concentration maps are available online (Connecticut DEP, 2006b)—including the 

2005 map shown in Exhibit 3-33. 

References 

Bricker, S.B., C.G. Clement, D.E. Pirhalla, S.P. Orlando, and D.R.G. Farrow. 1999. National 

eutrophication assessment: effects of nutrient enrichment in the nation’s estuaries. Silver Spring, MD: 

NOAA National Ocean Service. 71 pp. <http://ian.umces.edu/neea/pdfs/eutro_report.pdf> 

CAST (Council for Agricultural Science and Technology). 1999. Gulf of Mexico hypoxia: land and sea 

interactions. Task force report no. 134. 

CENR (Committee on Environment and Natural Resources). 2000. Integrated assessment of hypoxia in 

the northern Gulf of Mexico. Washington, DC: National Science and Technology Council Committee on 

Environment and Natural Resources. <http://www.nos.noaa.gov/Products/pubs_hypox.html> 

Connecticut DEP (Department of Environmental Protection). 2006a. DEP monitors Long Island Sound. 

Accessed 2006. <http://dep.state.ct.us/wtr/lis/monitoring/monsum.htm> 

Connecticut DEP (Department of Environmental Protection). 2006b. The Long Island Sound Water 

Quality Monitoring Program. Accessed 2006. <http://dep.state.ct.us/wtr/lis/monitoring/lis_page.htm> 

Howell, P., and D. Simpson. 1994. Abundance of marine resources in relation to dissolved oxygen in 

Long Island Sound. Estuaries 17:394-402. 

Jickells, T.D. 1998. Nutrient biogeochemistry of the coastal zone. Science 281:217-221. 

LUMCON (Louisiana Universities Marine Consortium). 2006. Hypoxia in the northern Gulf of Mexico. 

Accessed January 2006. <http://www.gulfhypoxia.net/> 

NRC (National Research Council). 2000. Clean coastal waters: understanding and reducing the effects of 

nutrient pollution. Washington, DC: National Academy Press. 405 pp. 

Rabalais, N.N., and R.E. Turner, eds. 2001. Coastal hypoxia: consequences for living resources and 

ecosystems. Coastal and estuarine studies 58. Washington, DC: American Geophysical Union. 454 pp. 

U.S. Commission on Ocean Policy. 2004. An ocean blueprint for the 21st century. Final report. 

Washington, DC. 

U.S. EPA. 2000. Ambient aquatic life water quality criteria for dissolved oxygen (saltwater): Cape Cod to 

Cape Hatteras. EPA/822/R-00/12. 

Vitousek, P.M., et al. 1997. Human alteration of the global nitrogen cycle: sources and consequences, 

ecological applications. Ecol. Appl. 7(3):737-750. 

Welsh, B.L., and F.C. Eller. 1991. Mechanisms controlling summertime oxygen depletion in western 

Long Island Sound. Estuaries 14:265-278. 

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INDICATOR: Harmful Algal Bloom Outbreaks Along the Western Florida 

Coastline

Harmful algal blooms (HABs) are “blooms” of large numbers of microscopic algae (phytoplankton) that 

can harm humans or the environment, especially when they occur in near-shore coastal waters and 

estuaries. HABs can be caused by a number of different species of phytoplankton. For example, in the 

Gulf of Mexico, approximately 50 species of toxic or potentially toxic marine microalgae have been 

identified (Fisher et al., 2003). HAB events along the Gulf coast are most commonly caused by the 

phytoplankton organism Karenia brevis (a dinoflagellate), also known as “red tide.” K. brevis can cause 

massive fish kills, marine mammal mortality, and in humans can cause neurotoxic shellfish poisoning 

(NSP) and respiratory irritation (NRC, 2000). 

HABs can occur naturally, but some may be exacerbated by excess nutrients from terrestrial activities 

(NRC, 2000). In the Gulf of Mexico, red tide events occur almost every year, generally in late summer or 

early fall. The extent and duration of these blooms largely reflect the influence of winds, currents, and 

other factors such as ocean mixing parameters. Red tide events in the Gulf typically last three to six 

months and cover hundreds of square miles, although in extreme cases, blooms can last up to 18 months 

and cover thousands of square miles (Haverkamp et al., 2004). K. brevis blooms tend to be concentrated 

along the west coast of Florida and, to a lesser extent, along the Texas coast (HABSOS, 2004). These 

events can impact ecological systems, fishing and shellfishing, and recreational activities in the Gulf of 

Mexico. For example, severe red tide blooms in 1996 resulted in fish mortalities and beach and shellfish 

bed closures throughout the Gulf, and killed over 150 endangered manatees along the Florida coast 

(HABSOS, 2002). 

This indicator reports the occurrence of elevated K. brevis concentrations along Florida’s Gulf Coast, 

where the Florida Fish and Wildlife Research Institute (FWRI) oversees an extensive red tide monitoring 

network. FWRI and its partners (Mote Marine Laboratory, the University of South Florida, Collier 

County, the Red Tide Offshore Monitoring Program, and others) maintain a set of permanent monitoring 

sites from which they collect water samples at regular intervals for microscopic examination. When a 

bloom is detected, researchers conduct additional targeted sampling to characterize the event. These 

efforts are aided by NOAA satellite imagery, which is used to track the movement of blooms and to direct 

targeted sampling. 

For this indicator, K. brevis concentrations from FWRI’s database were analyzed to determine the 

probability of a sample exceeding 5,000 K. brevis cells per liter in any given month (Christman and 

Young, 2006). At this concentration, the harvesting of shellfish in the area is prohibited (ISSC, 1999). For 

reference, background levels of K. brevis in the Gulf of Mexico are approximately 1,000 cells per liter 

year-round (Geesy and Tester, 1993). Although FWRI’s database extends back as far as the 1950s, a 

statistical review determined that prior to 2001, samples were not collected with sufficient regularity to 

provide meaningful information about trends over time (Christman and Young, 2006). Therefore, this 

indicator is restricted to the period 2001-2005. The analysis also was limited to a specific geographic area 

with a high sampling density (map in Exhibit 3-34). 

What the Data Show

Over the period of record (2001-2005), the probability of finding a K. brevis concentration above 5,000 

cells per liter in a sample taken from the study area ranged from less than 5 percent to more than 70 

percent, with both extremes occurring within the span of a few months in late 2001 and 2002 (Exhibit 3-

34). The graphic also suggests a seasonal pattern, with K. brevis counts peaking in late fall or early 
winter; this pattern is particularly well-defined in 

2001 and 2002. There is no discernable trend from 

year to year, however, and this period of record is 

currently too short to provide information about 

longer-term trends. 

Indicator Limitations 

• These data are biased toward surface and 

inshore sampling. The data do not 

include blooms occurring well offshore. 

• This indicator does not include HABs 

other than K. brevis. 

Data Sources 

Probability values were provided by the authors of a 

statistical review of the Florida Fish and Wildlife 

Research Institute (FWRI) database (Christman and 

Young, 2006). This database can be accessed by 

contacting FWRI (http://ocean.floridamarine.org/). 

References 

Christman, M., and L. Young. 2006. Analysis of 

Karenia Brevis Gulf Data. Report to the Florida Fish 

and Wildlife Research Institute. 

<http://research.myfwc.com/features/view_article.asp 

?id=27095> (executive summary only) 

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Fisher, W.S., T.C. Malone, and J.D. Giattina. 2003. A pilot project to detect and forecast harmful algal 

blooms in the northern Gulf of Mexico. Environ. Monit. Assess. 81(1-3):373-381. 

Geesy, M., and P.A. Tester. 1993. Gymnodinium breve: ubiquitous in the Gulf of Mexico waters? In: 

Smayda, T.J., and Y. Shimizu, eds. Toxic phytoplankton blooms in the sea. New York: Elsevier. pp. 251-

255. 

HABSOS. 2004. An integrated case study for the Gulf of Mexico. Final report. 

HABSOS. 2002. Prospectus: the Gulf of Mexico pilot project for a harmful algal blooms observing 

system (HABSOS). 

Haverkamp, D., K.A. Steidinger, and C.A. Heil. 2004. HAB monitoring and databases: the Florida 

Karenia brevis example. In: Hall, S., S. Etheridge, D. Anderson, J. Kleindinst, M. Zhu, and Y. Zou, eds. 

Harmful algae management and mitigation. APEC publication #204-MR-04.2. Singapore: Asia-Pacific 

Economic Cooperation. pp. 102-109. 

ISSC (Interstate Shellfish Sanitation Conference). 1999. National shellfish sanitation program: model 

ordinance. <http://issc.org/Archives/1999%20Reports/1999-Archive-Information.htm> 

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NRC (National Research Council). 2000. Clean coastal waters: understanding and reducing the effects of 

nutrient pollution. Washington, DC: National Academy Press. 405 pp. 

<http://www.nap.edu/books/0309069483/html/> 

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3.5.3 Discussion 

What These Indicators Say About Trends in the Extent and Condition of 

Coastal Waters and Their Effects on Human Health and the Environment 

Extent 

Although the ROE indicators do not characterize the extent of coastal waters, the Wetlands indicator (p. 

3-53) shows that at least one type of coastal system has experienced changes in extent over the last halfcentury. The number of acres of marine and estuarine wetlands has decreased overall since the 1950s, 

although the rate of loss has slowed in recent years. While the indicator does not identify the exact 

stressors responsible for the decline in marine and estuarine wetlands, it does list several factors that have 

led to overall wetland loss, including development and conversion to deepwater. Section 3.4 provides 

further detail on how human activities can affect wetland extent, including human activities that 

exacerbate natural processes (e.g., storm damage). Ultimately, trends in wetland extent affect ecological 

systems, as described further below. 

Condition 

Together, these indicators cover much of the spectrum of “condition,” including four of the broad themes 

introduced in Section 3.5.1: nutrients, toxic chemical contaminants, biotoxins, and the condition of native 

populations and their habitat. As described in Section 3.5.1, excess nutrients can cause algal blooms that 

result in low dissolved oxygen and reduced water clarity, which in turn can harm plant and animal 

communities. For example, the Trophic State of Coastal Waters indicator (p. 3-62) shows elevated levels 

of nutrients and chlorophyll-a (a surrogate for algal abundance) in a small but substantial portion of the 

nation’s estuarine areas. These results are consistent with indicators that show evidence of eutrophication, 

such as decreased water clarity and hypoxia. The SAV in Chesapeake Bay indicator (p. 3-74) in turn 

offers an example of an ecological effect linked to eutrophication. Nutrient stressors cannot be attributed 

entirely to human activities; for example, the Gulf of Mexico hypoxic zone results in part from natural 

mixing parameters, and trends in the extent of hypoxic zones show large year-to-year variations related to 

factors like climate (Hypoxia in Gulf of Mexico and Long Island Sound indicator, p. 3-77). However, as 

the spatial distribution of hypoxia in Long Island Sound suggests, the nation’s coastal waters can 

experience eutrophic effects that are very closely related to human activities (e.g., the location of a large 

city). Further, as the SAV in Chesapeake Bay indicator (p. 3-74) shows, present conditions may be quite 

different from historical reference conditions. 

Overall, levels of toxic chemical contaminants are low in most of the nation’s estuarine sediments, but as 

the Coastal Sediment Quality indicator (p. 3-67) shows, condition can vary greatly from one region to the 

next. In some EPA Regions, as much as 20 percent of estuarine area has sediments that either exceed 

contamination reference standards or fail a screening test for benthic toxicity. Other indicators discuss the 

extent to which toxic contaminants may be entering and affecting the food web. For example, benthic 

communities—which are most directly impacted by contaminants in sediment—generally show little 

evidence of disturbance (e.g., losses of pollution-sensitive species) (Coastal Benthic Communities 

indicator, p. 3-71). However, fish tissues had at least one contaminant above human health guidelines in 

22 percent of estuarine sampling sites (Coastal Fish Tissue indicator, p. 3-103), suggesting that 

bioaccumulation of certain toxic compounds is widespread and, in some instances, could pose risks to 

human health. This indicator suggests the importance of atmospheric deposition of mercury as a stressor 

to coastal water condition, as well as historical activities that released PCBs and DDT into upstream and 

coastal waters. 

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The HAB Outbreaks in Western Florida indicator (p. 3-81) describes the pervasive nature of red tide, 

which is one of many marine organisms that can produce dangerous biotoxins. As the data show, it is 

common for red tide events in the eastern Gulf of Mexico to persist for two or more months of the year at 

levels that trigger concerns about shellfish poisoning and associated human health effects (5,000 cells per 

liter). In extreme cases, these events also can be harmful to ecosystems. As this indicator shows, the 

condition of coastal waters with respect to HABs can vary greatly from year to year, depending on a 

number of factors. 

In more general ecological terms (populations, communities, and habitat), trends in the condition of 

coastal waters vary. Benthic communities in the nation’s estuaries are largely intact in terms of species 

diversity (Coastal Benthic Communities indicator, p. 3-71), which is critical because these organisms are 

a fundamental link in the coastal food web. Other populations, however, may be substantially lower than 

historical levels as a result of human stressors—for example, the Chesapeake Bay’s SAV, which is 

vulnerable to changes in water clarity (SAV in Chesapeake Bay indicator, p. 3-74). SAV is ecologically 

important because it is not just a plant population; it also provides habitat and facilitates nutrient cycling, 

much like wetlands do. SAV has recently shown increases in extent, which may translate into increased 

habitat and breeding grounds for various species. However, coastal habitat still continues to be threatened 

by human stressors. As the Hypoxia in Gulf of Mexico and Long Island Sound indicator (p. 3-77) shows, 

large areas of some of the nation’s coastal water bodies are unsuitable for fish and shellfish populations 

for at least a portion of the year. 

Limitations, Gaps, and Challenges 

Although the seven indicators discussed here provide a good overview of many important aspects of 

coastal extent and condition, there are a few key limitations to their temporal and spatial coverage. For 

example, the four indicators derived from the National Coastal Condition Report do not provide 

information about trends over time, as there are insufficient data from previous surveys to compare with 

recent data to examine potential trends.24 Another temporal limitation is that many surveys are conducted 

during an index period, not over a full year; thus, they may not capture phenomena that occur outside the 

sampling window.25 Spatially, the National Indicators are limited because they do not include data from 

Alaska, Hawaii, and most U.S. territories. Alaska contains 75 percent of the bays, sounds, and estuarine 

surface area in the United States, while Hawaii, the Caribbean, and the Pacific territories represent a set of 

unique estuarine subsystems (i.e., coral reefs and tropical bays) that are not common in the contiguous 48 

states. 

One challenge in assessing coastal waters is that some aspects of condition vary naturally from one area 

to another. For example, some rivers naturally carry a heavy load of sediments or nutrients into coastal 

waters, while benthic community structure may depend on climate, depth, and geology. To assess coastal 

waters with respect to natural background conditions, several of the ROE indicators use different 

reference conditions for different regions. 

24 U.S. EPA. 2004. National coastal condition report II. EPA/620/R-03/002. Washington, DC. 

<http://www.epa.gov/owow/oceans/nccr/2005/index.html> 

25 U.S. EPA. 2004. National coastal condition report II. EPA/620/R-03/002. Washington, DC. 

<http://www.epa.gov/owow/oceans/nccr/2005/index.html> 

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To assess the extent and condition of coastal waters more fully, it would help to have more information in 

several key areas, including: 

• More information about the extent of coastal waters—e.g., an indicator on coastal subsidence. 

• Nationally consistent data on coastal water pollutants beyond those associated with trophic 

state—for example, organics, toxics, metals, and pathogens. 

• A National Indicator of invasive species, which are often transported from one area to another 

along shipping routes or via aquaculture. Little information exists on a national level, in part 

because of a lack of standard invasion metrics. 

• Comprehensive information on the condition of the nation’s coral reefs—a unique and fragile 

habitat—and the status of coastal fish and shellfish communities.26 

26 U.S. EPA. 2004. National coastal condition report II. EPA/620/R-03/002. Washington, DC. 

<http://www.epa.gov/owow/oceans/nccr/2005/downloads.html> 

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3.6 WHAT ARE THE TRENDS IN THE QUALITY OF DRINKING WATER AND 

THEIR EFFECTS ON HUMAN HEALTH? 

3.6.1 Introduction 

The average American consumes 1-2 liters of drinking water per day, including water used to make 

coffee, tea, and other beverages.27 Virtually all drinking water in the United States comes from fresh 

surface water and ground water. Large-scale water supply systems tend to rely on surface water resources 

such as lakes, rivers, and reservoirs; these include the systems serving many large metropolitan areas. 

Smaller systems are more likely to use ground water, particularly in regions with limited surface water 

resources. Slightly more than half of the nation’s population receives its drinking water from ground 

water; i.e., through wells drilled into aquifers28 (including private wells serving about 15 percent of U.S. 

households29). If drinking water contains unsafe levels of contaminants, this contaminated water can 

cause a range of adverse human health effects. Among the potential effects are gastrointestinal illnesses, 

nervous system or reproductive effects, and chronic diseases such as cancer. 

Surface waters and aquifers can be contaminated by various agents, including microbial agents such as 

viruses, bacteria, or parasites (e.g., E. coli, Cryptosporidium, or Giardia); chemical contaminants such as 

inorganic metals, volatile organic compounds (VOCs), and other natural or manmade compounds; and 

radionuclides, which may be manmade or naturally occurring. Contaminants also can enter drinking water 

between the treatment plant and the tap (for example, lead can leach into water from old plumbing 

fixtures or household or street-side pipes). 

Drinking water contaminants can come from many sources: 

• Human activities that contaminate the source. Aquifers and surface waters that provide 

drinking water can be contaminated by many sources, as discussed in Sections 3.2 and 3.3. 

For example, chemicals from disposal sites or underground storage facilities can migrate into 

aquifers; possible contaminants include organic solvents (e.g., some VOCs), petroleum 

products, and heavy metals. Contaminants can also enter ground water or surface water as a 

result of their application to the land. Pesticides and fertilizer compounds (e.g., nitrate) can be 

carried into lakes and streams by rainfall runoff or snowmelt, or percolate through the ground 

and enter aquifers. Industrial wastes can contaminate drinking water sources if injected into 

27 U.S. EPA. 1997. Exposure factors handbook. Volume I—general factors. EPA/600/P-95/002Fa. Washington, DC. 

August. 

28 U.S. Geological Survey. 1999. Ground water (general interest publication). Reston, VA. 

<http://capp.water.usgs.gov/GIP/gw_gip/> 

29 U.S. EPA. 2002. The clean water and drinking water infrastructure gap analysis. EPA/816/R-02/020. 

<http://www.epa.gov/safewater/gapreport.pdf> 

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containment wells or discharged into surface waters, as can mine waste (e.g., heavy metals) if 

not properly contained. 

• Natural sources. As ground water travels through rock and soil, it can pick up naturally 

occurring contaminants such as arsenic, other heavy metals, or radionuclides. Some aquifers 

are naturally unsuitable for drinking because the local geology happens to include high levels 

of certain contaminants. 

• Microbial pathogens. Human wastes from sewage and septic systems can carry harmful 

microbes into drinking water sources, as can wastes from animal feedlots and wildlife. Major 

contaminants include Giardia, Cryptosporidium, and E. coli O157:H7. Coliform bacteria 

from human and animal wastes also may be found in drinking water if the water is not 

properly finished; these bacteria may indicate that other harmful pathogens are present as 

well. 

• Treatment and distribution. While treatment can remove many chemical and biological 

contaminants from the water, it may also result in the presence of certain disinfection 

byproducts that may themselves be harmful, such as trihalomethanes. Finished water can also 

become contaminated after it enters the distribution system, either from a breach in the 

system or from corrosion of plumbing materials, particularly those containing lead or copper. 

After water leaves the treatment plant, monitoring for lead in drinking water is done at the 

tap, and monitoring for microbial contaminants (as well as disinfection byproducts) occurs 

within the distribution system. 

Chemical exposure through drinking water can lead to a variety of long- and short-term effects. Potential 

health effects of exposure to certain metals, solvents, and pesticides can include chronic conditions such 

as cancer, which can develop over long periods of time (up to 70 years). Higher doses over shorter 

periods of time can result in a variety of biological responses, including toxicity, mutagenicity, and 

teratogenicity (birth defects). Short-term results might include cosmetic effects (e.g., skin discoloration), 

unpleasant odors, or more severe problems such as nervous system or organ damage, and developmental 

or reproductive effects. The effects of some drinking water contaminants are not yet well understood. For 

example, certain disinfection byproducts have been associated with cancer, developmental, and 

reproductive risks, but the extent of this association is still uncertain. 

Consuming water with pathogenic microbes can cause life-threatening diseases such as typhoid fever or 

cholera—rare in the U.S. today—as well as more common waterborne diseases caused by organisms such 

as Giardia, Cryptosporidium, E. coli, and Campylobacter. Health consequences of the more common 

illnesses can include symptoms such as gastrointestinal distress (stomach pain, vomiting, diarrhea), 

headache, fever, and kidney failure, as well as various infectious diseases such as hepatitis. 

A number of factors determine whether the presence of contaminants in drinking water will lead to 

adverse health effects. These include the type of contaminant, its concentration in the water, individual 

susceptibility, the amount of contaminated water consumed, and the duration of exposure. 

Disinfection of drinking water—the destruction of pathogens using chlorine or other chemicals—has 

dramatically reduced the incidence of waterborne diseases such as typhoid, cholera, and hepatitis, as well 

as gastrointestinal illness, in the United States. Other processes required depend on the physical, 

microbiological, and chemical characteristics and the types of contaminants present in the source water 

(e.g., filtration to remove turbidity and biological contaminants; treatment to remove organic chemicals 

and inorganic contaminants such as metals; and corrosion control to reduce the presence of corrosion 

byproducts such as lead at the point of use). 

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3.6.2 ROE Indicators 

This section presents an indicator that tracks trends in the total population served by community water 

systems (CWS) for which states report no violations of health-based drinking water standards. Data for 

this indicator come from EPA’s Safe Drinking Water Information System, Federal Version. This system 

houses all data submitted by states, EPA Regions, and the Navajo Nation Indian Tribe on the community 

water systems they oversee. 

Table 3.6.1. ROE Indicators of the Trends in the Quality of Drinking Water and their Effects on 

Human Health 

NATIONAL INDICATORS LOCATION 

Population Served by Community Water Systems with No Reported 

Violations of Health-Based Standards (N/R) 

3.6.2 – p. 3-90 

N/R = National Indicator displayed at EPA Regional scale 

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INDICATOR: Population Served by Community Water Systems with No Reported 

Violations of Health-Based Standards 

Community Water Systems (CWS), public water systems that supply water to the same population yearround, served over 281 million Americans in fiscal year (FY) 2005 (U.S. EPA, 2006)—more than 90 

percent of the U.S. population (U.S. Census Bureau, 2005). This indicator presents the percentage of 

Americans served by CWS for which states reported no violations of EPA health-based standards for over 

90 contaminants (U.S. EPA, 2004b). 

Health-based standards include Maximum Contaminant Levels (MCLs) and Treatment Techniques (TTs). 

An MCL is the highest level of a contaminant that is allowed in drinking water. A TT is a required 

treatment process (such as filtration or disinfection) intended to prevent the occurrence of a contaminant 

in drinking water (U.S. EPA, 2004c). TTs are adopted where it is not economically or technologically 

feasible to ascertain the level of a contaminant, such as microbes, where even single organisms that occur 

unpredictably or episodically can cause adverse health effects. Compliance with TTs may require finished 

water sampling, along with quantitative or descriptive measurements of process performance to gauge the 

efficacy of the treatment process. MCL-regulated contaminants tend to have long-term rather than acute 

health effects, and concentrations vary seasonally (if at all; e.g., levels of naturally occurring chemical 

contaminants or radionuclides in ground water are relatively constant). Thus, compliance is based on 

averages of seasonal, annual, or less frequent sampling. 

This indicator tracks the population served by CWS for which no violations were reported to EPA for the 

period from FY 1993 to FY 2005, the latest year for which data are available. Results are reported as a 

percentage of the overall population served by CWS, both nationally and by EPA Region. This indicator 

also reports the number of persons served by systems with reported violations of standards covering 

surface water treatment, microbial contaminants (microorganisms that can cause disease), and 

disinfection byproducts (chemicals that that may form when disinfectants, such as chlorine, react with 

naturally occurring materials in water and may pose health risks) (U.S. EPA, 2004b). The indicator is 

based on violations reported quarterly by States, EPA, and the Navajo Nation Indian Tribe, who each 

review monitoring results for the CWS that they oversee. 

What the Data Show

Of the population served by CWS nationally, the percentage served by systems for which no health-based 

violations were reported for the entire year increased from 79 percent in 1993 to 94 percent in FY 2002 

before declining to 89 percent in FY 2005 (Exhibit 3-35). This indicator is based on reported violations of 

the standards in effect in any given year. Several new standards went into effect after December 31, 2001. 

These were the first new drinking water standards to take effect during the period of record (beginning in 

1993). The results after FY 2002 would have been somewhat higher had it not been for violations of 

standards that became effective in FY 2002 or after (Exhibit 3-35; see dark segment atop the last three 

columns). As EPA adds to or strengthens its requirements for water systems over time, compliance with 

standards comes to represent a higher level of public health protection. 

When results are broken down by EPA Region, some variability over time is evident (Exhibit 3-36). 

Between FY 1993 and FY 2005, most Regions were consistently above the national percentage. Three of 

the Regions were substantially below the national average over much of the period of record, but as of FY 

2005, only one Region remained well below the national percentage, largely because of a small number of 

public water systems serving large populations. 
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In FY 2005, reported violations involving surface water treatment rules in large CWS were responsible 

for exceeding health-based standards for more than 14 million people (5 percent of the population served 

by CWS nationally) (Exhibit 3-37). Reported violations of heath-based coliform standards affected 11.6 

million people (4.1 percent of the CWS-served population), and reported violations of the health-based 

disinfection byproducts standards (Stage 1) affected nearly 6 million people (2.1 percent of the CWSserved population). Overall, of the 11.5 percent of the population served by systems with reported 

violations in FY 2005, 85 percent of these cases involved at least one of these three rules governing 

treatment to prevent waterborne diseases—the most widespread and acute threat to health from drinking 

water—or the contaminants created by such treatment. 

Indicator Limitations 

• Non-community water systems (typically relatively small systems) that serve only transient 

populations such as restaurants or campgrounds, or serving those in a non-domestic setting 

for only part of their day (e.g., a school, religious facility, or office building), are not included 

in population served figures. 

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• Domestic (home) use of drinking water 

supplied by private wells—which serve 

approximately 15 percent of the U.S. 

population (USGS, 2004)—is not 

included. 

• Bottled water, which is regulated by 

standards set by the Food and Drug 

Administration, is not included. 

• National statistics based on population 

served can be volatile, because a single 

very large system can sway the results by 

up to 2 to 3 percent; this effect becomes 

more pronounced when statistics are 

broken down at the regional level, and 

still more so for a single rule. 

• Some factors may lead to overstating the 

extent of population receiving water that 

violates standards. For example, the 

entire population served by each system 

in violation is reported, even though only 

part of the total population served may 

actually receive water that is out of 

compliance. In addition, violations stated 

on an annual basis may suggest a longer 

duration of violation than may be the 

case, as some violations may be as brief 

as an hour or a day. 

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• Other factors may lead to understating the population receiving water that violates standards. 

CWS that purchase water from other CWS are not always required to sample for all 

contaminants themselves, and the CWS that are wholesale sellers of water generally do not 

report violations for the population served by the systems that purchase the water. 

• Under-reporting and late reporting of violations by states to EPA affect the ability to 

accurately report the national violations total. EPA estimated that between 1999 and 2001, 

states were not reporting 35 percent of all health-based violations, which reflects a sharp 

improvement in the quality of violations data compared to the previous three-year period 

(U.S. EPA, 2004a). In late 2006, EPA expects to issue an updated estimate of data quality for 

the period 2002-2004. 

• State data verification and other quality assurance analyses indicate that the most widespread 

data quality problem is under-reporting of monitoring and health-based violations and 

inventory characteristics. Under-reporting occurs most frequently in monitoring violations; 

even though these are separate from the health-based violations covered by the indicator, 

failures to monitor could mask violations of TTs and MCLs. 

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Data Sources 

Data for this indicator were obtained from EPA’s Safe Drinking Water Information System (U.S. EPA, 

2006) (http://www.epa.gov/safewater/data/getdata.html; 

http://www.epa.gov/safewater/data/pivottables.html). This database contains a record of violations 

reported to EPA by the states or other entities that oversee Community Water Systems, along with annual 

summary statistics. 

References 

U.S. Census Bureau. 2005. Annual population estimates 2000 to 2005. Released December 22, 2005. 

Washington, DC. <http://www.census.gov/popest/states/tables/NST-EST2005-01.xls> Available from 

<http://www.census.gov/popest/states/NST-ann-est.html> 

U.S EPA. 2006. Safe Drinking Water Information System, Federal Version. Accessed July 2006. 

<http://www.epa.gov/safewater/data/getdata.html> 

U.S. EPA. 2004a. Safe Drinking Water Act 30th anniversary fact sheet: drinking water monitoring, 

compliance, and enforcement. 

<http://www.epa.gov/safewater/sdwa/30th/factsheets/monitoring_compliance.html> 

U.S. EPA. 2004b. Safe Drinking Water Act 30th anniversary fact sheet: drinking water standards and 

health effects. <http://www.epa.gov/safewater/sdwa/30th/factsheets/standard.html> 

U.S. EPA. 2004c. Safe Drinking Water Act 30th anniversary fact sheet: glossary. 

<http://www.epa.gov/safewater/sdwa/30th/factsheets/glossary.html> 

USGS (U.S. Geological Survey). 2004. Estimated use of water in the United States in 2000. 2004 

revision. <http://water.usgs.gov/pubs/circ/2004/circ1268/> 

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3.6.3 Discussion 

What This Indicator Says About Trends in the Quality of Drinking Water and 

Their Effects on Human Health 

Most Americans served by CWS are served by facilities with no reported violations (Drinking Water 

indicator, p. 3-90). Since 1993, the percentage of Americans served by community water systems for 

which states reported no health-based violations has increased, although there has been some reversal 

nationally since the percentage peaked in 2002. While there have been noticeable differences among EPA 

Regions over the period of record, most Regions have been consistently above 90 percent since 1993. 

Only one Region has been consistently below the national average, though according to the data source, 

this result is due largely to one large metropolitan water system which is under a legal settlement to 

upgrade its treatment technology. As this result suggests, while the nation has thousands of community 

water systems, a substantial percentage of the population depends on the quality of a small number of 

large metropolitan water systems. 

Limitations, Gaps, and Challenges 

As noted in the indicator description, a challenge in assessing national drinking water quality is that there 

are inherent limitations in using reporting data. Some violations may be unreported, particularly if 

monitoring is inadequate—leading to undercounting. Other violations may be overlooked because CWS 

may purchase water from other CWS and not test it for all contaminants themselves. Conversely, the data 

could also overstate the portion of the population receiving water in violation of standards, because a 

violation could be as short as an hour or a day and be limited to water received by only a small portion of 

a system’s customers. 

Other challenges relate to the interpretation of the Drinking Water indicator (p. 3-90). For example, trends 

can be confounded by the fact that water quality standards and treatment requirements change over time. 

Thus, an apparent increase in violations over time may result from new or more stringent MCLs rather 

than simply a decline in the quality of drinking water, as these new requirements may also affect some 

systems’ compliance with existing standards. 

As described in the indicator summary, the indicator does not address the quality of drinking water other 

than that obtained from CWSs. Information that would provide a more complete characterization of 

drinking water quality include National Indicators for: 

• Trends in drinking water quality from CWS that did have reported violations. The 

Drinking Water indicator does not explain the nature of every reported violation; nor does it 

show how many contaminants may be above standards, the identity of the contaminants, the 

extent to which standards were exceeded, or the duration of the violations (some of which, 

especially in larger systems, were only a very few hours in length). 

• The quality of drinking water from other public water systems. There is no ROE 

indicator for drinking water quality from transient and non-transient non-community water 

systems, which are required to monitor quality and report violations to state authorities, but 

are regulated only for certain contaminants. 

• The quality of drinking water from non-public water supplies. Private wells, cisterns, and 

other non-public water supplies are not subject to federal regulation. Some private supplies 

are treated, and some people do test their private water for common contaminants. However, 

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no national infrastructure, and few if any systematic state efforts, currently exist to collect 

data on trends in the quality of these supplies. Bottled water is regulated by the Food and 

Drug Administration (FDA), which is required by law to apply standards that are no less 

stringent or protective of public health than EPA’s, but there is no ROE indicator on the 

quality of bottled water. 

In addition to these gaps, there are no ROE indicators to identify trends in health effects of interest, such 

as waterborne disease occurrence. Data are very limited for endemic waterborne illness as well as for 

acute waterborne disease outbreaks. 

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3.7 WHAT ARE THE TRENDS IN THE CONDITION OF RECREATIONAL WATERS 

AND THEIR EFFECTS ON HUMAN HEALTH AND THE ENVIRONMENT? 

3.7.1 Introduction 

The nation’s rivers, lakes, and coastal waters are used for many different forms of recreation. Some 

recreational activities take place in or on the water, such as swimming, boating, whitewater rafting, and 

surfing. Other activities may not involve contact with the water yet may still require water—or be 

enhanced by proximity to water. Examples include a picnic at the beach, hiking, nature viewing (e.g., bird 

watching), and hunting (especially waterfowl). People also engage in fishing and shellfishing as 

recreational activities. 

In the questions on fresh surface waters and coastal waters (Sections 3.2 and 3.5), condition is defined as 

a combination of physical, chemical, and biological attributes of a water body. For recreational waters, 

condition is more specific, focusing on those physical, chemical, and biological attributes that determine a 

water body’s ability to support recreational activities. The particular attributes necessary to support 

recreation vary widely, depending on the nature of the activity in question. In a more general sense, 

however, the components of recreational condition fall into two main categories: 

• Attributes that determine whether recreational activities can be enjoyed without unacceptable 

risk to human health—primarily pathogens and chemical contaminants that can affect the 

health of humans who are exposed during contact activities such as swimming. 

• Attributes associated with ecological systems that support recreation—e.g., the status of fish 

and bird communities, as well as chemical and physical characteristics that may affect these 

populations and their habitat. These attributes also contribute to the aesthetic qualities 

important for recreational activities. 

Many stressors affecting the condition of recreational waters fall into the broad category of contaminants. 

This category includes chemical contaminants, various pathogens (viruses, bacteria, and other parasites or 

protozoans) that can cause infectious disease, and pollutants such as trash or debris. These stressors can 

come from a variety of point sources and nonpoint sources, and can be discharged or washed directly into 

recreational waters or carried downstream to lakes or coastal areas. Among the major sources are storm 

water and sediment runoff, direct discharge (e.g., from industrial facilities and sewer systems), 

atmospheric deposition, and recreational activities themselves (e.g., outboard motor exhaust and 

overboard discharge of sanitary wastes). Some chemicals and pathogens occur naturally, but their 

abundance may be influenced by other human stressors such as land use and land cover (e.g., paved 

surfaces and forestry and irrigation practices, which can influence runoff patterns) or by natural stressors 

such as weather and climate. Land use and land cover can influence recreational condition in other ways 

as well. 

In terms of human health, the stressors that pose the greatest potential risks are chemical and biological 

contaminants. People can be exposed to these contaminants if they swim in contaminated waters or near 

storm water or sewage outfall pipes—especially after a rainfall event. Boating also may pose risks of 

exposure, although to a lesser extent. For toxic chemical contaminants, the main routes of exposure are 

through dermal (skin) contact or accidental ingestion. For pathogens, the main route of exposure is by 

swallowing water, although some infections can be contracted simply by getting polluted water on the 

skin or in the eyes. In some cases, swimmers can develop illnesses or infections if an open wound is 

exposed to contaminated water. 

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Effects of exposure to chemical and biological contaminants range from minor illnesses to potentially 

fatal diseases. The most common illness is gastroenteritis, an inflammation of the stomach and the 

intestines that can cause symptoms such as vomiting, headaches, and diarrhea. Other minor illnesses 

include ear, eye, nose, and throat infections. While unpleasant, most swimming-related illnesses are 

indeed minor, with no long-term effects. However, in severely contaminated waters, swimmers can 

sometimes be exposed to serious and potentially fatal diseases such as meningitis, encephalitis, hepatitis, 

cholera, and typhoid fever.30 Children, the elderly, and people with weakened immune systems are most 

likely to develop illnesses or infections after coming into contact with contaminated water. 

From an ecological perspective, stressors to recreational waters can affect habitat, species composition, 

and important ecological processes. For example, changes in land cover (e.g., the removal of shade trees) 

may cause water temperature to rise above the viable range for certain fish species. Hydromodifications 

such as dams may create some recreational opportunities (e.g., boating), but they also may impede the 

migration of fish species such as salmon. Chemical and biological contaminants may harm plants and 

animals directly, or they may disrupt the balance of the food web. For example, acid deposition may lead 

to acidification in lakes, while excess nutrients can lead to eutrophic conditions such as low levels of 

dissolved oxygen, which in turn can harm fish and shellfish populations. Beyond their obvious effects on 

activities like fishing and nature viewing, stressors such as these also can be detrimental to recreational 

activities in a more aesthetic sense, as the presence of dead fish or visibly unhealthy plants may diminish 

one’s enjoyment of recreation in or near the water. 

Ultimately, ecological effects can also impact human health. For example, eutrophic conditions can 

encourage harmful algal blooms (HAB)—some of which can produce discomfort or illness when people 

are exposed through ingestion or skin or eye contact. One well-known type of HAB is “red tide,” which in 

humans can cause neurotoxic shellfish poisoning (NSP) and respiratory irritation.

31 

3.7.2 ROE Indicators 

At this time, no National Indicators have been identified to quantify the condition of recreational waters. 

Individual states monitor certain recreational waters for a set of indicator bacteria and report monitoring 

results to EPA. However, the methodology and frequency of data collection vary among states, so the 

data are not necessarily comparable. 

Challenges and information gaps for developing reliable National Indicators of recreational water 

condition are described in more detail in Section 3.7.3 below. 

30 Pond, K. 2005. Water recreation and disease—plausibility of associations, sequelae and mortality. Published on 

behalf of World Heath Organization. London, United Kingdom: IWA Publishers. 

<http://www.who.int/water_sanitation_health/bathing/recreadis.pdf> 

31 National Research Council. 2000. Clean coastal waters: understanding and reducing the effects of nutrient 

pollution. Washington, DC: National Academy Press. 405 pp. 

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3.7.3 Discussion 

Limitations, Gaps, and Challenges 

Several challenges exist in assessing the condition of the nation’s recreational waters. Foremost is the lack 

of a comprehensive national system for collecting data on pathogen levels at beaches, a key concern in 

assessing the suitability of recreational waters with respect to human health. In addition, data on the types 

and extent of health effects associated with swimming in contaminated water are limited. The number of 

occurrences is likely under-reported because individuals may not link common symptoms (e.g., 

gastrointestinal ailments, sore throats) to exposure to contaminated recreational waters. 

Another challenge to answering this question is the breadth of the subject. “Recreation” encompasses a 

wide range of activities, involving different types of water bodies and entailing varying concepts of 

condition. While the recreational condition of a whitewater stream with a native salmon population will 

be determined largely by flow levels and condition of fish habitat, for example, the recreational condition 

of a beach will be assessed more in terms of levels of pathogens and chemical contaminants. 

Gaps in assessing the condition of the nation’s recreational waters include National Indicators of pathogen 

levels in recreational waters (rivers, lakes, and coastal beaches), the magnitude of specific stressors— 

particularly contaminant loadings (biological and chemical)—to recreational waters, harmful algal 

blooms (HABs) in recreational waters, and the condition of recreational fish and shellfish populations. 

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3.8 WHAT ARE THE TRENDS IN THE CONDITION OF CONSUMABLE FISH AND 

SHELLFISH AND THEIR EFFECTS ON HUMAN HEALTH? 

3.8.1 Introduction 

Fish and shellfish caught through commercial, recreational, or subsistence fishing are an important part of 

a healthful diet for many people. Fish and shellfish contain high-quality protein and other essential 

nutrients, are low in saturated fat, and contain omega-3 fatty acids. Most fish consumed in the United 

States is commercial fish, purchased in supermarkets or fish stores. Fishing also is one of the most 

popular outdoor recreational activities in the country, with more than 34 million people per year fishing 

recreationally32—many of whom eat at least some of the fish they catch. In addition, subsistence fishers— 

people who rely on fish as an affordable food source or for whom fish are culturally important—consume 

fish and shellfish as a major part of their diets. Commercial, recreational, and subsistence fisheries all 

have substantial economic value for the nation, regions, and local communities. 

Americans consume fish and shellfish caught in the nation’s lakes, rivers, and estuaries and in deep ocean 

fisheries, as well as farmed fish and shellfish. Some of these fish and shellfish contain elevated levels of 

chemical or biological contaminants. This question addresses the condition of consumable fish and 

shellfish caught or farmed in the United States—whether, and the extent to which, these organisms 

contain contaminants that could affect the health of people who consume them. 

According to recent surveys, the average American consumes close to 13 grams of fish and shellfish per 

day (prepared weight), which amounts to slightly more than one 3-ounce serving per week.33 However, 

many Americans consume substantially more fish and shellfish than the national average; some of the 

highest consumption rates are among tribal and ethnic populations who fish for subsistence. Concern 

about fish and shellfish safety is higher for these groups as well as for children, pregnant and nursing 

women (because of possible effects on the fetus or infant), and other population subgroups who may be 

more vulnerable to the health effects of certain chemical or biological contaminants (e.g., elderly or 

immunosuppressed individuals). 

Chemical contaminants of greatest concern in consumable fish and shellfish tend to be those that are 

persistent, bioaccumulative, and toxic (called PBTs). These chemicals can persist for long periods in 

sediments and then enter the food web when ingested by bottom-dwelling (benthic) organisms. Benthic 

organisms are eaten by smaller fish, which in turn are eaten by larger fish, which may be consumed by 

humans or wildlife. PBTs that are common in fresh and coastal waters include: 

• Mercury. This highly toxic metal is present in waters all over the globe—a result of longrange transport and deposition of airborne mercury as well as direct inputs to water.34 

Mercury in water bodies can be methylated by certain bacteria in bottom sediments to form 

32 U.S. Department of the Interior, Fish and Wildlife Service, and U.S. Department of Commerce, U.S. Census 

Bureau. 2002. 2001 national survey of fishing, hunting, and wildlife-associated recreation. 

33 U.S. EPA. 2002. Estimated per capita fish consumption in the United States. EPA/821/C-02/003. Washington, 

DC. <http://www.epa.gov/waterscience/fish/consumption_report.pdf> 

34 U.S. and global sources of mercury are described in more detail in Section 2.2, which includes an indicator of 

domestic mercury emissions. 

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methylmercury, which is more toxic and bioavailable than other forms of mercury.

35 It also is 

biomagnified through aquatic food webs, so that it becomes particularly concentrated in 

larger and longer-lived predators such as bass, tuna, swordfish, and some sharks. Exposure to 

high levels of methylmercury can cause reproductive and other effects in wildlife;36 in 

humans, exposure to elevated levels is primarily associated with developmental and 

neurological health effects.37 

• Polychlorinated biphenyls (PCBs) and the pesticide DDT. Though PCBs and DDT are no 

longer manufactured or used in the U.S., they persist in historical deposits in watersheds and 

near-shore sediments, which can continue to contaminate fish and shellfish. These chemicals 

are also circulated globally as a result of use in other parts of the world. Levels of PCBs and 

DDT are a concern in some bottom-feeding fish and shellfish, as well as in some higher-level 

predators. These chemicals have been linked to adverse health effects such as cancer, nervous 

system damage, reproductive disorders, and disruption of the immune system in both humans 

and wildlife. 

Other chemical contaminants that may be present in fish and shellfish include other pesticides, metals 

(such as arsenic), and dioxins and furans.

38 

Biological contamination also can affect the condition of fish and shellfish—particularly the latter. For 

example, shellfish contaminated with pathogens from human and animal fecal wastes can cause 

gastrointestinal illness and even death in individuals with compromised immune systems. Sources of fecal 

contamination in shellfish include urban runoff, wildlife, wastewater treatment systems and treatment 

plants, agricultural runoff, and boating and marinas. 

Marine biotoxins produced by certain types of algae can contaminate fish and shellfish as well. These 

toxins not only can harm fish and fish communities—sometimes resulting in massive fish kills or losses 

to aquaculture operations—but they also can make their way through the food web to affect seabirds, 

marine mammals, and humans. Mollusks such as mussels, clams, oysters, whelks, and other shellfish can 

carry biotoxins that have common symptoms such as irritation of the eyes, nose, throat, and tingling of 

the lips and tongue. Consumption of contaminated seafood can cause a range of other health effects in 

humans, depending on the organism involved, including gastrointestinal illness, amnesia, memory loss, 

paralysis, and even death.39,40 

35 U.S. EPA. 1997. Mercury study report to Congress. Volume III: fate and transport of mercury in the environment. 

EPA/452/R-97/005. <http://www.epa.gov/mercury/report.htm> 

36 U.S. EPA. 1997. Mercury study report to Congress. Volume V: health effects of mercury and mercury 

compounds. EPA/452/R-97/007. <http://www.epa.gov/mercury/report.htm> 

37 National Research Council. 2000. Toxicological effects of methylmercury. Washington, DC: National Academies 

Press. 

38 U.S. EPA. In progress. National study of chemical residues in lake fish tissue. Washington, DC. 

<http://www.epa.gov/waterscience/fishstudy> 

39 Baden D., L.E. Fleming, and J.A. Bean. 1995. Marine toxins. In: DeWolff, F.A., ed. Handbook of clinical 

neurology: intoxications of the nervous system, part II: natural toxins and drugs. Amsterdam, The Netherlands: 

Elsevier Press. pp. 141-175. 

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The growth of aquaculture, or fish farming, may affect the levels of certain contaminants in consumable 

fish and shellfish. Dense colonies can increase stress and disease transmission among fish, in some cases 

requiring the administration of antibiotics.41 Studies have also found higher levels of certain contaminants 

in farmed fish than in their wild counterparts, possibly due to differences in diet. For example, several 

studies have found higher concentrations of PCBs, organochlorine pesticides, and polybrominated 

diphenyl ethers (PBDEs) in farmed salmon.42 

Overfishing also can affect the condition of fish—not only the species being fished, but also the species 

that prey on them—by disrupting the food web. Because of depleted food sources, predators can become 

more susceptible to disease (such as infection of rockfish by mycobacterial lesions). These infections are 

often confined to internal organs and may not be apparent to anglers, although in some cases they are 

associated with external sores as well. Some types of mycobacteria can also infect humans who handle 

diseased fish if the infection comes into contact with an open wound. The slow-developing infections are 

usually not severe in humans, but in some cases they can cause major health problems, especially in 

people with compromised immune systems. 

3.8.2 ROE Indicators 

Two ROE indicators characterize levels of chemical contaminants in edible fish and shellfish species. 

One indicator reports levels and occurrence of contaminants in fish in estuarine areas; the other, in 

freshwater lakes and reservoirs. Both indicators are based on nationwide probabilistic surveys. 

The coastal fish indicator is based on an index originally presented in EPA’s second National Coastal 

Condition Report (NCCR II). The underlying data were collected between 1997 and 2000 as part of 

EPA’s Environmental Monitoring and Assessment Program (EMAP). EMAP’s probabilistic coastal 

surveys are designed to be representative of 100 percent of estuarine acreage in the contiguous 48 states. 

This indicator presents results by EPA Region. 

The other indicator describes contamination of fish in inland lakes. This indicator is derived from fish 

samples collected and analyzed for EPA’s National Study of Chemical Residues in Lake Fish Tissue, a 

probabilistic survey designed to estimate the national distribution of the mean levels of selected 

persistent, bioaccumulative, and toxic chemical residues in fish tissue from lakes and reservoirs. 

Note that this question does not rely on information about fish and shellfish consumption advisories. 

While many states and tribes issue fish consumption advice and develop fish advisory programs, there is 

great variability in how monitoring is conducted, how decisions are made to place waters under advisory, 

and what specific advice is provided when contamination is found in fish. Further, trends in the number of 

40 Van Dolah, F.M. 2000. Marine algal toxins: origins, health effects, and their increased occurrence. Environ. 

Health Persp. 108(Suppl 1):133-141. 

41 Barton, B.A., et al. 1991. Physiological changes in fish from stress in aquaculture with emphasis of the response 

and effects of corticosteroids. Annu. Rev. Fish Dis. 1:3-26. 

42 Easton, M.D.L., D. Luszniak, and E. Von der Geest. 2002. Preliminary examination of contaminant loadings in 

farmed salmon, wild salmon and commercial salmon feed. Chemosphere 46(7):1053-1074. 

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advisories over time may reflect changes in the frequency and intensity of monitoring.43 Thus, fish 

advisories cannot provide a consistent national metric for trends in the condition of consumable fish and 

shellfish. 

Table 3.8.1. ROE Indicators of the Trends in the Condition of Consumable Fish and Shellfish and 

their Effects on Human Health 

NATIONAL INDICATORS LOCATION 

Coastal Fish Tissue Contaminants (N/R) 3.8.2– p. 3-103 

Contaminants in Lake Fish Tissue 3.8.2– p. 3-107 

N/R = National Indicator displayed at EPA Regional scale 

43 U.S. EPA. 2005. Fact sheet: national listing of fish advisories. EPA/823/F-05/004. 

<http://epa.gov/waterscience/fish/advisories/fs2004.pdf> 

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1 INDICATOR: Coastal Fish Tissue Contaminants 

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Contaminants in fish not only affect their own health and ability to reproduce, but also affect the many 

species that feed on them. Contaminants also may make fish unsuitable for human consumption (U.S. 

EPA, 2000).

This indicator, derived from an indicator presented in EPA’s second National Coastal Condition Report 

(NCCR II) (U.S. EPA, 2004), is based on National Coastal Assessment (NCA) fish tissue survey data 

from 653 estuarine sites throughout the United States. The survey was designed to provide a national 

picture of coastal fish tissue contaminants by sampling sites in estuarine waters throughout the contiguous 

48 states. Each site was sampled once during the period 1997-2000, within an index period from July to 

September. The indicator reflects average condition during this index period. 

Fish and shellfish analyzed in the survey included Atlantic croaker, white perch, catfish, flounder, scup, 

blue crab, lobster, shrimp, whiffs, mullet, tomcod, spot, weakfish, halibut, sole, sculpins, sanddabs, bass, 

and sturgeon. At each site, five to ten whole-body fish samples were tested for 90 contaminants, 16 of 

which have EPA-established risk guidelines for recreational fishers. This indicator is based on data 

collected from 1997 to 2000. 

To assess risks to human health, contaminant concentrations in fish tissue were compared with 

established EPA guidelines based on the consumption of four 8-ounce fish meals per month (U.S. EPA, 

2000, 2004). For most contaminants this is done using whole body concentrations, but for mercury, which 

concentrates in the edible fillet portion of the fish, a factor of 3.0 was used to correct whole-body 

concentrations in order to approximate fillet concentrations. The factor, 3.0, represents the median value 

(range 1.5-5.0) found in the available literature (Windom and Kendall, 1979; Mikac et al., 1985; Schmidt 

and Brumbaugh, 1990; Kannan et al., 1998; Canadian Council of Ministers of the Environment, 1999). 

For this indicator, a site was given a high score if one or more contaminants were present at a 

concentration above the guideline ranges. A site was rated moderate if one or more contaminants were 

within the guideline ranges but none was in exceedance. Sites with all contaminants below their guideline 

ranges were given a low contamination score. 

What the Data Show

Nationwide, 63 percent of sites showed low fish tissue contamination, 15 percent had moderate 

contamination, and 22 percent exhibited high contamination (Exhibit 3-38). Fish tissue contamination 

varied notably from one EPA Region to the next; for example, the percentage of sites with low 

contamination ranged from 25 percent (Region 1) to 83 percent (Region 4). Regions 2 and 9 had the 

largest proportion of sites with high contamination (41 percent and 40 percent, respectively). 

Data from EPA’s EMAP National Coastal Database showed that nationwide, PCBs were the 

contaminants most frequently responsible for high fish tissue contamination, with 19 percent of sites 

above EPA guidelines (Exhibit 3-39). Other chemicals present above EPA guidelines at many sites were 

mercury in muscle tissue (18 percent of sites), DDT (8 percent), and PAHs (3 percent) (Exhibit 3-39). 

Inorganic arsenic, selenium, chlordane, endosulfan, endrin, heptachlor epoxide, hexachlorobenzene, 

lindane, and Mirex were below EPA guidelines for all fish sampled in the NCA. 
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Indicator Limitations 

• The indicator does not include data from Louisiana, Florida, Puerto Rico, Alaska, and 

Hawaii, which had not been assessed at the time this indicator was compiled. Some of these 

areas (e.g., portions of Alaska) have now been surveyed, and may be included in future 

indicators. 

• Whole-body contaminant concentrations in fish overestimate the risk associated with 

consuming only the fillet portion of the fish, with the exception of mercury and cadmium, 

which are generally underestimated. 

• This indicator focuses on contaminants from a human health risk perspective. No EPA 

guidance criteria exist to assess the ecological risk of whole-body contaminants in fish (U.S. 

EPA, 2004). 

• Some fish samples used in the survey were non-market-size juveniles, which are known to 

have lower contaminant levels than larger, market-sized fish. 

• Samples are collected during an index period from July to September, and the indicator is 

only representative of this time period. It is unlikely, however, that contaminant levels vary 

substantially from season to season. 

• There are no trend data for this indicator. In NCCR II, fish tissue contaminants are 

characterized by whole-body concentrations and compared to EPA risk-based consumption 

guidelines. For EPA’s first National Coastal Condition Report (NCCR I), fish contaminants 

were measured as fillet concentrations and compared to FDA criteria. The data presented here 

will serve as a baseline for future surveys, however. 

Data Sources 

This indicator is based on an analysis published in EPA’s National Coastal Condition Report II (U.S. 

EPA, 2004). Summary data by EPA Region and by contaminant have not been published, but were 

provided by EPA’s National Coastal Assessment program. Underlying sampling data are housed in 

EPA’s National Coastal Assessment database (U.S. EPA, 2005) 

(http://www.epa.gov/emap/nca/html/data/index.html). 

References 

Canadian Council of Ministers of the Environment. 1999. Protocol for the derivation of Canadian tissue 

residue guidelines for the protection of wildlife that consume aquatic biota. Prepared by the Task Force 

on Water Quality Guidelines. 

Kannan, K., R.G. Smith, R.F. Lee, H.L. Windom, P.T. Heimuller, J.M. Macauley, and J.K. Summers. 

1998. Distribution of total mercury and methyl mercury in water, sediment and fish from South Florida 

estuaries. Arch. Environ. Con. Tox. 34:109-118. 

Mikac, N., M. Picer, P. Stegnar, and M. Tusek-Nidari. 1985. Mercury distribution in a polluted marine 

area, ratio of total mercury, methyl mercury and selenium in sediments, mussels and fish. Water Res. 

19:1387-1392. 

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Schmidt, C.J., and W.G. Brumbaugh. 1990. National contaminant biomonitoring program: concentrations 

of arsenic, cadmium, copper, lead, mercury selenium and zinc in U.S. freshwater fish 1976-1984. Arch. 

Environ. Con. Toxicol. 19:731-747. 

U.S. EPA. 2005. EMAP national coastal database. Accessed 2005. 

<http://www.epa.gov/emap/nca/html/data/index.html> 

U.S. EPA. 2004. National coastal condition report II. EPA/620/R-03/002. Washington, DC. 

<http://www.epa.gov/owow/oceans/nccr/2005/index.html> 

U.S. EPA. 2000. Guidance for assessing chemical contaminant data for use in fish advisories. 

EPA/823/B-00/008. <http://www.epa.gov/waterscience/library/fish/> 

Windom, H.L., and D.R. Kendall. 1979. Accumulation and biotransformation of mercury in coastal and 

marine biota. In: Nriagu, J.O., ed. Biogeochemistry of mercury in the environment. Amsterdam, The 

Netherlands: Elsevier. pp. 303-323. 

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1 INDICATOR: Contaminants in Lake Fish Tissue

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Lakes and reservoirs provide important sport fisheries and other recreational opportunities, and lake 

ecosystems provide critical habitat for aquatic species and support wildlife populations that depend on 

aquatic species for food. Lakes and reservoirs occur in a variety of landscapes and can receive 

contaminants from several sources, including direct discharges into the water, atmospheric deposition, 

and agricultural or urban runoff. A group of contaminants of particular concern are the persistent, 

bioaccumulative, and toxic (PBT) chemicals. These contaminants are highly toxic, long-lasting chemicals 

that can accumulate in fish, reaching levels that can affect the health of people and wildlife that eat them. 

PBT contaminants can originate from a variety of sources, many of which are declining. The primary 

source of one of the most important PBTs, mercury, is atmospheric deposition. Among other important 

PBTs, most uses of DDT became illegal in the U.S. effective in 1973; production of PCBs in the U.S. 

ceased in 1977 (although they are still emitted as a byproduct of other manufacturing processes) and most 

uses phased out in 1979; chlordane was banned in 1988; and quantifiable emissions of dioxin-like 

compounds from all known sources have decreased in the U.S. by an estimated 89 percent between 1987 

and 2000 (U.S. EPA, 2006a). 

This indicator is based on tissue samples of predator and bottom-dwelling fish species collected and 

analyzed for EPA’s National Study of Chemical Residues in Lake Fish Tissue. The data generated from 

this probabilistic survey (Olsen et al., 1998; U.S. EPA, 1999; Stevens and Olsen, 2003, 2004) are 

designed to estimate the national distribution of the mean levels of PBT chemicals in fish tissue from 

lakes and reservoirs of the contiguous 48 states. Fish samples were collected from 500 lakes and 

reservoirs over a four-year period (2000-2003). Sampling locations were selected from the estimated 

147,000 target lakes and reservoirs in the contiguous 48 states based on an unequal probability survey 

design. The lakes and reservoirs were divided into six size categories, and varying probabilities were 

assigned to each category in order to achieve a similar number of lakes in each size category. The lakes 

and reservoirs ranged from 1 hectare (about 2.5 acres) to 365,000 hectares (about 900,000 acres), were at 

least 1 meter (3 feet) deep, and had permanent fish populations. 

Because no predator or bottom-dwelling species occurs in all 500 lakes and reservoirs, the study focused 

on 12 target predator species and 6 target bottom-dwelling species in order to minimize the effect of 

sampling different species. These species were chosen because they are commonly consumed in the study 

area, have a wide geographic distribution, and potentially accumulate high concentrations of PBT 

chemicals. Sampling teams applied consistent materials and methods nationwide. From each lake or 

reservoir, teams collected composite samples of five adult fish of similar size for one predator species 

(e.g., bass or trout) and one bottom-dwelling species (e.g., carp or catfish) (U.S. EPA, 2000). Fillets were 

analyzed for predators, and whole bodies were analyzed for bottom-dwelling fish. Fillet data represent the 

edible part of the fish most relevant to human health, while whole body data are more relevant to wildlife 

consumption. A single laboratory prepared fish tissue samples for analysis in a strictly controlled 

environment, and tissue samples were sent to four analytical laboratories. The same laboratory analyzed 

tissue samples for each chemical group (e.g., PCBs or organochlorine pesticides), using the same standard 

analytical method, for the duration of the study. The indicator consists of statistical distributions of the 

concentrations of 15 PBT chemicals or chemical groups in predator and bottom-dwelling fish tissue, 

including mercury, arsenic (total inorganic), dioxins/furans, total PCBs, and 11 organochlorine pesticides. 

Fourteen of these chemicals or chemical groups also appear in the Coastal Fish Tissue indicator (p. 3-

103).
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What the Data Show 

Mercury, polychlorinated biphenyls (PCBs), dioxins and furans, and DDT are widely distributed in lakes 

and reservoirs in the contiguous 48 states (Exhibits 3-40, 3-41). Mercury and PCBs were detected in 100 

percent of both predator and bottom-dweller composite samples. Dioxins and furans were detected in 81 

percent of the predator composite samples and 99 percent of the bottom-dweller composite samples, and 

DDT was detected in 78 percent of the predator composites and 98 percent of the bottom-dweller 

composites. Some of the chemicals analyzed for this study were not detected in any of the fish tissue 

samples. 

Median concentrations in predator fillets (i.e., half of the lakes and reservoirs had fish with higher values) 

were as follows: mercury, 0.285 ppm; total PCBs, 2.161 ppb; dioxins and furans, 0.006 ppt [TEQ]; and 

total DDT, 1.47 ppb (Exhibit 3-40). Median concentrations in whole, bottom-dwelling fish were lower for 

mercury (0.069 ppm), but higher for total PCBs (13.88 ppb), dioxins and furans (0.406 ppt [TEQ]), and 

total DDT (12.68 ppb) (Exhibit 3-41). 

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Indicator Limitations 

• Survey data are not available for Alaska, Hawaii, and Puerto Rico. 

• The Great Lakes, the Great Salt Lake, and lakes without permanent fish populations are not 

included in the target population. 

• Because the distribution of sampling sites was based on the frequency of occurrence of lakes 

and reservoirs, contaminants in lakes and reservoirs in arid states (e.g., Arizona, New 

Mexico, and Nevada) are not well-represented. 

• Due to the inaccessibility (e.g., landowner denial of access) of some target lakes, the results 

are representative of the sampled population of lakes (approximately 80,000) rather than the 

original target population of 147,000 lakes. 

• Trend data are not yet available, as this is the first time that a national lake fish tissue survey 

has been conducted using a probabilistic sampling design. These data will serve as a baseline 

for future surveys. 

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Data Sources 

The data for Exhibits 3-40 and 3-41 were provided by EPA’s National Lake Fish Tissue Study. A report 

on the findings of this study was still in progress at the time this Report on the Environment went to 

press; however, partial results (number of detections) have been published in U.S. EPA (2006b) 

(http://www.epa.gov/waterscience/fishstudy/results.htm), along with information about how to obtain 

more detailed results on CD. 

References 

Olsen, A.R., D.L. Stevens, Jr., and D. White. 1998. Application of global grids in environmental 

sampling. Comp. Sci. Stat. 30:279-284. 

Stevens, D.L., Jr., and A.R. Olsen. 2004. Spatially-balanced sampling of natural resources. J. Am. Stat. 

Assoc. 99(465):262-278. 

Stevens, D.L., Jr., and A.R. Olsen. 2003. Variance estimation for spatially balanced samples of 

environmental resources. Environmetrics 14:593-610. 

U.S. EPA. 2006a. Inventory of sources of environmental releases of dioxin-like compounds in the United 

States: the year 2000 update. EPA/600/P-03/002a. Washington, DC. [link will be provided at future date] 

U.S. EPA. 2006b. National Lake Fish Tissue Study—results. Updated February 27, 2006. 

<http://www.epa.gov/waterscience/fishstudy/results.htm> 

U.S. EPA. 2000. Field sampling plan for the national study of chemical residues in lake fish tissue. 

EPA/823/R-02/004. Washington, DC. 

U.S. EPA. 1999. National study of chemical residues in lake fish tissue: study design. Washington, DC. 

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3.8.3 Discussion 

What These Indicators Say About Trends in the Condition of Consumable 

Fish and Shellfish and Their Effects on Human Health 

These indicators provide baseline information about consumable fish in inland lakes, reservoirs, and 

coastal areas. The data were collected from a variety of species, reflecting many parts of the food web. 

The results for fish in estuarine sites along the Atlantic, Gulf, and Pacific coasts of the contiguous 48 

states (Coastal Fish Tissue indicator, p. 3-103) varied substantially among the seven coastal EPA 

Regions. Fish from the coastal waters of the Southeast (EPA Region 4) generally had “low” 

contamination scores, while several other Regions had a substantial proportion with “high” 

contamination. In general, PCBs, mercury, and PAHs appeared to be the contaminants responsible for the 

most “high” contamination scores. 

The results for lake fish (Lake Fish Tissue indicator, p. 3-107) suggest that several chemical contaminants 

are widely distributed in the nation’s lakes and reservoirs, including mercury, dioxins and furans, PCBs, 

and DDT. However, some of the other chemicals in this screening—including certain pesticides and 

PAHs—were detected rarely or not at all. There were some notable differences between predators and 

bottom-dwellers, which may be a result of how each type of fish was analyzed—fillets for predators and 

whole fish for bottom dwellers. 

Limitations, Gaps, and Challenges 

As explained in Section 3.8.2, both of these indicators have important limitations. For example, like the 

other coastal indicators from NCCR II (presented in Section 3.5), the Coastal Fish Tissue indicator (p. 3-

103) does not display trend data. It is also limited spatially, as adequate data for Alaska, Hawaii, the 

Caribbean, and the Pacific territories are not available. The lack of data from Alaska is especially notable 

because more than half of the nation’s commercial fish and shellfish catch comes from Alaskan waters.44 

The Lake Fish Tissue indicator (p. 3-107) is also limited temporally and spatially, with no trend data and 

no coverage outside the contiguous 48 states. Further, unlike the coastal survey, the lake fish survey was 

not designed to produce results by region, and it also does not compare contaminant levels to any healthbased guidelines. Thus, while both indicators present meaningful data, the results cannot easily be 

compared. 

The Lake Fish Tissue and Coastal Fish Tissue indicators (pp. 3-107 and 3-103) do provide some 

information about contamination and safety of fish and shellfish. However, to fully assess the condition of 

the nation’s fish and shellfish, more data are needed—particularly on a national level, because many 

issues have been studied locally or regionally, but have not yet been studied in nationally representative 

surveys. In addition to the limitations of the indicators described above, information gaps for answering 

this question include nationally consistent indicators of pathogens in fish and shellfish, in both fresh water 

and coastal waters, and of the biological and chemical condition of fish and shellfish commercially 

farmed in the U.S. There are also no ROE indicators to describe the effects of fish and shellfish condition 

on human health. As noted in Chapter 1, it is often difficult to explicitly connect an observed effect to a 

44 National Oceanic and Atmospheric Administration. 2005. Fisheries of the United States—2004.

<http://www.st.nmfs.gov/st1/fus/fus04/fus_2004.pdf>

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particular stressor (e.g., the condition of fish and shellfish that people consume), even though there may 

be scientific evidence to suggest a possible association. 

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