U.S. Department of the Interior

U.S. Geological Survey

Circular 1291

The Quality of Our Nation’s Waters

National Water-Quality Assessment Program

Pesticides in the Nation’s Streams and 

Ground Water, 1992–2001
Front cover middle photograph © Corbis
The Quality of Our Nation’s Waters 

Pesticides in the Nation’s Streams and 

Ground Water, 1992–2001

By Robert J. Gilliom, Jack E. Barbash, Charles G. Crawford, Pixie A. Hamilton, 

Jeffrey D. Martin, Naomi Nakagaki, Lisa H. Nowell, Jonathan C. Scott, 

Paul E. Stackelberg, Gail P. Thelin, and David M. Wolock

Circular 1291

U.S. Department of the Interior

U.S. Geological Survey
U.S. Department of the Interior

DIRK KEMPTHORNE, Secretary

U.S. Geological Survey

P. Patrick Leahy, Acting Director

U.S. Geological Survey, Reston, Virginia: 2006

Revised February 15, 2007

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U.S. Government.

Although this report is in the public domain, permission must be secured from the individual copyright owners to 

reproduce any copyrighted materials contained within this report.

Suggested reference:

Gilliom and others, 2006, The Quality of Our Nation’s Waters—Pesticides in the Nation’s Streams and Ground Water, 

1992–2001: U.S. Geological Survey Circular 1291,172 p.

Library of Congress Cataloging-in-Publication Data

Pesticides in the nation’s streams and ground water, 1992–2001 : the quality of our nation’s waters / by Robert J. 

Gilliom ... [et al.].

 p. cm. -- (Circular 1291)

 Includes bibliographical references.

 1. Water--Pollution--United States. 2. Agricultural pollution--United States. 3. Pesticides--Environmental aspects. 4. 

Water quality--United States. I. Gilliom, Robert J. II. U.S. Geological Survey circular ; 1291. 

 TD428.A37P47 2006

 363.738’49820973--dc22

 2005037748

ISBN 1-411-30955-3
“The Heinz Center depends heavily on NAWQA data to support our 

periodic report: ‘The State of the Nation’s Ecosystems.’ NAWQA 

data provide the foundation for our description of chemical 

contamination—including pesticides and other compounds—both 

nationally and among different land uses, and for tracking how 

contaminant levels change over time. We appreciate NAWQA’s 

strong commitment to making its information and data readily 

accessible to meet our organization’s needs and to address the 

Nation’s water-resource information needs.” 

Robin O’Malley, Senior Fellow and Program Director, 

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

Environment 

“EPA has worked closely with the USGS NAWQA Program since 

it began in 1991, with the goal of advancing the scientific tools 

and data that are available to assess the potential risks posed by 

pesticides in surface water and ground water. Recently, EPA and 

USGS have collaborated in developing statistical models to predict 

concentrations of atrazine and other pesticides in streams around 

the country, including quantitative estimates of reliability. The models, 

developed by USGS from NAWQA data, increase EPA’s capacity to 

estimate surface-water concentrations of pesticides. Model results 

are being used to identify locations where additional monitoring may 

be needed to evaluate the ecological condition of watersheds.”

Elizabeth Behl, Chief of Environmental Risk Branch, 

Office of Pesticide Programs, 

U.S. Environmental Protection Agency

iii
NAWQA iv National Water-Quality Assessment Program

The U.S. Geological Survey (USGS) is committed to serving the Nation with accurate and timely 

scientific information that helps enhance and protect the overall quality of life, and facilitates effective 

management of water, biological, energy, and mineral resources (http://www.usgs.gov/). Information 

on the quality of the Nation’s water resources is of critical interest to the USGS because water 

quality is integrally linked to the long-term availability of water that is clean and safe for drinking 

and recreation and that is suitable for industry, irrigation, and habitat for fish and wildlife. Escalating 

population growth and increasing demands for multiple water uses make water availability, now 

measured in terms of quantity and quality, even more critical to the long-term sustainability of our 

communities and ecosystems.

The USGS implemented the National Water-Quality Assessment (NAWQA) Program (http://water.

usgs.gov/nawqa/) to support national, regional, and local information needs and decisions related 

to water-quality management and policy. Shaped by and coordinated with ongoing efforts of other 

Federal, State, and local agencies, the NAWQA Program is designed to answer: What is the quality 

of our Nation’s streams and ground water? How is the quality changing over time? How do natural 

features and human activities affect the quality of streams and ground water, and where are those 

effects most pronounced? By combining information on water chemistry, physical characteristics, 

stream habitat, and aquatic life, the NAWQA Program aims to provide science-based insights for 

current and emerging water issues and priorities. NAWQA results can contribute to informed decisions 

that result in practical and effective water-resource management and strategies that protect and 

restore water quality.

Since 1991, the NAWQA Program has implemented interdisciplinary assessments in 51 of the 

Nation’s most important river basins and aquifers, referred to as Study Units, and the High Plains 

Regional Ground Water Study (see accompanying map and list of studies). Collectively, these areas 

account for more than 70 percent of total water use (excluding thermoelectric and hydropower) and 

more than 50 percent of the population’s supply of drinking water. The areas are representative of the 

Nation’s major hydrologic landscapes, priority ecological resources, and agricultural, urban, and natural 

sources of contamination. 

Each assessment is guided by a nationally consistent study design and methods of sampling and 

analysis. The assessments thereby build local knowledge about water-quality issues and trends in a 

particular stream or aquifer while providing an understanding of how and why water quality varies 

regionally and nationally. The consistent, multi-scale approach helps to determine if certain types of 

water-quality issues are isolated or pervasive, and allows direct comparisons of how human activities 

and natural processes affect water quality and ecological health in the Nation’s diverse geographic 

and environmental settings. Comprehensive assessments of pesticides, nutrients, volatile organic 

compounds, trace elements, and aquatic ecology are developed at the national scale through national 

data analysis and comparative analysis of the Study Unit findings.

The USGS places high value on the communication and dissemination of credible, timely, and 

relevant science so that the most recent and available knowledge about water resources can be 

applied in management and policy decisions. We hope this NAWQA publication will provide you the 

needed insights and information to meet your needs, and thereby foster increased awareness and 

involvement in the protection and restoration of our Nation’s waters. 

The NAWQA Program recognizes that a national assessment by a single program cannot address 

all water-resource issues of interest. External coordination at all levels is critical for a fully integrated 

understanding of watersheds and for cost-effective management, regulation, and conservation of our 

Nation’s water resources. The Program, therefore, depends extensively on the advice, cooperation, 

and information from other Federal, State, interstate, Tribal, and local agencies, nongovernmental 

organizations, industry, academia, and other stakeholder groups. The assistance and suggestions of all 

are greatly appreciated.

 Robert M. Hirsch

Associate Director, Water
NAWQA Study Units

1 Acadian–Pontchartrain Drainages

2 Albemarle–Pamlico Drainage Basin

3 Allegheny and Monongahela River Basins

4 Apalachicola–Chattahoochee–Flint River Basin

5 Central Arizona Basins

6 Central Columbia Plateau

7 Central Nebraska Basins

8 Connecticut, Housatonic, and Thames River Basins

9 Cook Inlet Basin

10 Delaware River Basin

11 Delmarva Peninsula

12 Eastern Iowa Basins

13 Georgia–Florida Coastal Plain

14 Great and Little Miami River Basins

15 Great Salt Lake Basins

16 Hudson River Basin

17 Island of Oahu

18 Kanawha–New River Basins

19 Lake Erie–Lake Saint Clair Drainages

20 Long Island–New Jersey Coastal Drainages

21 Lower Illinois River Basin

22 Lower Susquehanna River Basin

23 Lower Tennessee River Basin

24 Las Vegas Valley Area and the Carson and Truckee River Basins

25 Mississippi Embayment

26 Mobile River Basin

27 New England Coastal Basins

28 Northern Rockies Intermontane Basins

29 Ozark Plateaus

30 Potomac River Basin

31 Puget Sound Basin

32 Red River of the North Basin

33 Rio Grande Valley

34 Sacramento River Basin

35 San Joaquin–Tulare Basins

36 Santa Ana Basin

37 Santee River Basin and Coastal Drainages

38 South-Central Texas

39 South Platte River Basin

40 Southern Florida

41 Trinity River Basin

42 Upper Colorado River Basin

43 Upper Illinois River Basin

44 Upper Mississippi River Basin

45 Upper Snake River Basin

46 Upper Tennessee River Basin

47 Western Lake Michigan Drainages

48 White River Basin

49 Willamette Basin

50 Yakima River Basin

51 Yellowstone River Basin

NAWQA Study Units

Initiated 1991

Initiated 1994

Initiated 1997

High Plains Regional Ground 

Water Study, initiated 1999

31

50

49

6 28

45 51

34 15

35

24

24

36

9

17

5

33

38

41

29

22 20

27

30

25

26

2

18

11

10

16

8

3

19

14 48

43

47

12

44

32

21

23

4 13

37

46

40

1

42

39

7

v
vi

Introduction to this report and the NAWQA series

The Quality of Our Nation’s Waters

This report is one of a series of publications, The Quality of Our Nation’s 

Waters, that describe major findings of the NAWQA Program on waterquality issues of regional and national concern. This report presents 

evaluations of pesticides in streams and ground water based on findings 

for the first decadal cycle of NAWQA. “Pesticides in the Nation’s Streams 

and Ground Water, 1992–2001” greatly expands the analysis of pesticides 

presented in “Nutrients and Pesticides,” which was the first report in the 

series and was based on early results from 1992 to 1995. Other reports in 

this series cover additional water-quality constituents of concern, such as 

volatile organic compounds and trace elements, as well as physical and 

chemical effects on aquatic ecosystems. Each report builds toward a more 

comprehensive understanding of regional and national water resources.

The information in this series is intended primarily for those interested 

or involved in resource management, conservation, regulation, and 

policymaking at regional and national levels. In addition, the information 

might interest those at a local level who wish to know more about the 

general quality of streams and ground water in areas near where they live 

and how that quality compares with other areas across the Nation.

 P. Patrick Leahy, Acting Director

 U.S. Geological Survey
Contents

Chapter 1 is a broad overview of key national findings and their implications. 

Chapters 2 and 3 provide background on pesticides and how they were assessed 

by NAWQA. Together, Chapters 4 and 5 provide a detailed assessment of findings 

regarding the occurrence and behavior of pesticides in streams and ground water. 

Chapter 6 evaluates the occurrence and distribution of pesticides in terms of the 

potential for effects on human health, aquatic life, and wildlife. Chapters 7 and 

8 examine two important topics with implications for the future—prediction of 

pesticide levels for unmonitored areas and emerging evidence of long-term trends. 

1 Overview of Findings and Implications .....................................................1

2 Pesticide Primer..........................................................................................21

3 NAWQA’s Approach to Pesticide Assessment......................................31

4 Occurrence and Distribution in Streams and Ground Water..............41

5 Complexities: Seasonality, Mixtures, and Degradates.........................67

6 Potential for Effects on Human Health, Aquatic Life, and Wildlife.....87

7 Prediction Where Data are Inadequate................................................119

8 Long-Term Trends .....................................................................................127

References Cited.......................................................................................137

Glossary......................................................................................................145

Appendix 1—Pesticide Compounds Analyzed ....................................153

Appendix 2—Properties Affecting Transport and Fate......................159

Appendix 3—Water-Quality Benchmarks ............................................161

Appendix 4—List of Abbreviations ........................................................171

vii

About 1 billion pounds of conventional pesticides are used 

each year in the United States to control weeds, insects, and 

other pests. The use of pesticides has resulted in a range of 

benefits, including increased food production and reduction 

of insect-borne disease, but also raises questions about pos-

sible adverse effects on the environment, including water 

quality. The NAWQA assessment of pesticides provides the 

most comprehensive national-scale analysis to date of pesticide 

occurrence and concentrations in streams and ground water. 

NAWQA results show where, when, and why specific pesti-

cides occur in streams and ground water across the Nation, and 

yield science-based implications for assessing and managing 

the quality of our water resources.

1

Overview of Findings and Implications

This chapter provides a broad 

overview of NAWQA findings 

about the occurrence and 

distribution of pesticides in the 

Nation’s streams and ground 

water and summarizes the 

implications of these findings 

for water-quality assessment 

and management. Priorities for 

filling remaining information 

gaps also are addressed. 

Detailed discussions of each 

major topic are provided in 

subsequent chapters, including 

selected case studies of 

pesticide occurrence within 

individual NAWQA Study Units.

Photograph by Rand Schaal
Introduction—

New results confirm and expand findings from 

earlier NAWQA studies

This report is based on the National WaterQuality Assessment (NAWQA) Program’s first 

decade of water-quality assessments, which were 

completed on a rotational schedule during 1992–

2001 in 51 major hydrologic systems across the 

country, referred to as Study Units (see p. iv and 

v). Assessments were conducted using a nationally consistent approach in 20 Study Units during 

1992–1995; in 16 Study Units during 1996–1998; 

and in 15 Study Units during 1998–2001.

Nationally, water samples for pesticide 

analysis were collected from 186 stream sites 

within the 51 Study Units, bed-sediment samples 

were collected from 1,052 stream sites, and fish 

samples were collected from 700 stream sites. 

Ground-water samples were collected from 

5,047 wells. In this report, most data analyses for 

stream water are based on the single year of most 

intensive sampling; data analyses for bed sediment and fish tissue are based on one composite 

sample per site; and data analyses for ground 

water are based on one sample per well. Sampling sites for streams and ground water were 

selected to represent the specific agricultural, 

urban, undeveloped, and mixed-land-use settings of greatest significance to water resources 

in the primary hydrologic settings within each of 

the Study Units. Shallow ground water (generally less than 20 ft below the water table) was 

sampled in agricultural, urban, and undeveloped 

areas, whereas deeper ground water was sampled 

from wells that tap major aquifers, most of which 

Relation to Previous Studies

Over the past 50 years, a vast amount of research has been conducted to 

investigate the spatial and temporal distributions of pesticides and their degradates in the hydrologic system, the biological effects of these compounds, 

and the myriad chemical, physical, and biological processes that control their 

transport and fate in the environment. Much of this previous work was summarized in a NAWQA book series entitled “Pesticides in the Hydrologic System,” 

which examined these issues in relation to pesticides in the atmosphere 

(Majewski and Capel, 1995), ground water (Barbash and Resek, 1996), surface 

water (Larson and others, 1997), and bed sediment and aquatic biota (Nowell 

and others, 1999). In addition, since this book series was published, there have 

been many more studies and new reviews of specific topics by scientists in 

government, academia, corporations, and other organizations. This report is not 

intended to be a comprehensive review of all of these topics, although investigations directly relevant to the findings discussed in this report are cited in the 

text. The focus of this report is on the summary and interpretation of NAWQA 

data collected during 1992–2001.

are affected by a mixture of land uses and are 

important as potential sources of drinking water. 

Most NAWQA water samples were analyzed 

for 75 pesticides and 8 pesticide degradates, 

including 20 of the 25 most heavily used herbicides and 16 of the 25 most heavily used insecticides, but few fungicides, fumigants, or other 

types of pesticides were analyzed. Degradates are 

new compounds formed by transformation of a 

pesticide by chemical, photochemical, or biological reactions. In addition, 32 organochlorine pesticide compounds were analyzed in bed sediment 

and (or) fish tissue—19 parent pesticides and 

13 degradates and manufacturing by-products 

(hereinafter referred to as by-products). Most of 

the organochlorine pesticides are no longer used 

in the United States, but the parent compounds, 

degradates, or by-products may persist in the 

environment. Pesticide compounds analyzed are 

listed in Appendix 1. 

This analysis of NAWQA results for 

1992–2001 builds upon an initial national assessment of pesticides in streams and ground water 

that was based on results from NAWQA’s first 

20 Study Unit investigations (summarized in the 

first report of this series, U.S. Geological Survey, 

1999). The more extensive data and expanded 

geographic coverage available for this report 

confirm and reinforce many of the previously 

reported findings, allow more detailed analyses 

of each topic, and support new analyses, such as 

the development of statistical models that extend 

the results from targeted NAWQA studies to 

areas of the Nation that have not been assessed. 

In addition, water-quality benchmarks for assessing the potential significance of pesticide concentrations to aquatic life and fish-eating wildlife 

have been substantially updated to incorporate 

the most recent values available from the U.S. 

Environmental Protection Agency (USEPA) and 

other sources.

NAWQA findings are summarized below for 

major topics, each of which is identified with the 

chapter in this report where more detailed results, 

explanations, and references are provided. Key 

implications are also summarized for each topic, 

focusing on the extension of study results to 

national assessment of water quality, applications to water-quality management, and needs for 

additional information.

The NAWQA approach and design are summarized in Chapter 3. Details on data-analysis 

methods, as well as all data used in this report, 

are available at http://ca.water.usgs.gov/pnsp/

pubs/circ1291/.

2 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Water-quality assessments by NAWQA, which is a single program 

among many local, State, and Federal programs, were not designed 

to address all of the Nation’s water-resource information needs and 

issues. Listed below are several characteristics and limitations of the 

NAWQA approach that are important to consider when interpreting 

the findings on pesticides presented in this report.

NAWQA assessments characterized the quality of the available, 

untreated water resources, and not the quality of drinking water 

(as would be done by monitoring water from water-treatment 

plants or from household taps). By focusing on the quality of 

streams and ground water in their present condition (ambient 

water quality), NAWQA complements many Federal, State, and 

local drinking-water monitoring programs.

NAWQA assessments did not focus on specific sites with known 

water-quality problems or narrowly defined “issues of the day,” 

but rather on the condition of the total resource, including 

streams and ground water in a wide range of hydrologic and 

land-use settings across the country. 

NAWQA assessments of pesticides focused primarily on nonpoint sources resulting from applications for pest management in 

agricultural, urban, and other land-use settings, although some 

sites—particularly those downstream from major metropolitan 

areas—also may be influenced by point sources, such as discharges from wastewater treatment plants.

NAWQA assessments targeted specific land-use settings that 

are most extensive or important to water quality in a wide range 

of hydrologic and environmental settings across the Nation. 

This targeted approach gives priority to understanding the most 

critical factors influencing water quality. Extension of results to 

national analysis, however, requires careful definition of each 

type of water resource and environmental setting for which conclusions are drawn and the use of statistical models to extrapolate results to resources that have not been measured.

USGS analytical methods were designed to measure concentrations as low as economically and technically feasible. Studies 

of contaminant occurrence and behavior benefit from the most 

information possible at all concentration levels, and such data 

help to identify emerging issues and to track changes in concentrations over time. By this approach, however, pesticides were 

commonly detected at concentrations far below Federal or State 

standards and guidelines for protecting water quality. Detections 

of pesticides do not necessarily indicate that there are appreciable risks to human health, aquatic life, or wildlife, which must 

be assessed by comparing measured concentrations with those 

that may cause adverse effects.

USGS methods for analyzing pesticides in water measured 

concentrations in filtered water samples and, thus, may underestimate concentrations of compounds that have strong affinities 

for suspended particles. The potential for underestimation is 

greater for stream water compared with ground water because 

of the generally greater amounts of suspended particles present 

in stream water—which are removed by filtration along with any 

pesticides contained in or on the particles.

•

•

•

•

•

•

Pesticide compounds analyzed in water by NAWQA included 

many of the most heavily used herbicides and insecticides, but 

they included only a fraction of all pesticides currently in use and 

few of their degradates. NAWQA findings provide insights about 

what to expect for pesticides and degradates that were not measured, but must be considered as only a partial assessment of 

currently used pesticides.

Organochlorine pesticide compounds analyzed by NAWQA in 

bed sediment and fish tissue are predominantly related to pesticides that were no longer in use by 1990. Of the pesticide compounds measured in bed sediment and fish tissue, only dacthal, 

endosulfan, lindane, methoxychlor, and permethrin were used 

during all or part of the study period.

•

•

NAWQA studies used nationally consistent methods for 

sample collection and laboratory analysis. Urban groundwater studies, for example, often required the installation 

of new observation wells to ensure comparable data 

among studies.

Chapter 1

Overview of Findings and Implications 3

Unique Features of the NAWQA Approach
Pesticides in water

(most were used during the study period)

Organochlorine compounds in fish and sediment

(most are no longer used)

Percentage of time (streams) or samples 

(ground water) with one or more detections 

Percentage of samples with

one or more detections

1_1freqexOVsummary

Shallow ground water

Shallow ground water

Stream water

Shallow ground water

Major aquifers

Agricultural areas

Urban areas

Undeveloped areas

Mixed land uses

Agricultural areas

Urban areas

Undeveloped areas

Mixed land uses

Stream water

Stream water

Stream water

61%

55%

 65%

29%

33%

97%

97%

94%

Overview of pesticide occurrence

25 50 75 100

Bed sediment

Bed sediment

Fish tissue

Bed sediment

Bed sediment

Fish tissue

Fish tissue

Fish tissue

57%

80%

 57%

24%

60%

92%

94%

94%

0 0 25 50 75 100

Figure 1–1. One or more pesticides or degradates were detected in water more than 90 percent of the 

time during the year in streams draining watersheds with agricultural, urban, and mixed land uses. In 

addition, some organochlorine pesticides that have not been used in the United States for many years 

were detected along with their degradates and by-products in most samples of whole fish or bed sediment 

from streams sampled in these land-use settings. Pesticides were less common in ground water, but were 

detected in more than 50 percent of wells sampled to assess shallow ground water in agricultural and urban 

areas.

Pesticide Occurrence—

Pesticides were frequently detected in streams 

and ground water (Chapter 4)

Pesticides or their degradates were detected 

in one or more water samples from every stream 

sampled. One or more pesticides or degradates 

were detected in water more than 90 percent of 

the time during the year in agricultural streams, 

urban streams, and mixed-land-use streams 

(fig. 1–1). This finding is based on a timeweighted analysis of results for 4,380 water samples, which adjusts results for variable sampling 

frequencies to avoid biases that may be caused by 

differences in sampling intensity among sites and 

seasons. Undeveloped streams had one or more 

detectable pesticides or degradates 65 percent of 

the time. The presence of pesticide compounds 

in predominantly undeveloped watersheds may 

result from past or present uses within the watershed for purposes such as forest management or 

maintenance of rights-of-way, uses associated 

with small areas of urban or agricultural land, or 

atmospheric transport from other areas. 

Organochlorine pesticides (such as DDT) 

and their degradates and by-products were 

found in fish or bed-sediment samples from 

most streams in agricultural, urban, and mixedland-use settings—and in more than half the 

fish samples from streams draining undeveloped 

watersheds (fig. 1–1). Most organochlorine pesticides had not been used in the United States for a 

number of years prior to the study period, but the 

continued occurrence of some historically used 

organochlorine pesticide compounds demonstrates their persistence in the environment. 

Pesticides were less common in ground 

water than in streams (fig. 1–1). Nevertheless, 

more than half of the shallow wells sampled in 

agricultural and urban areas, and 33 percent of 

the deeper wells that tap major aquifers, which 

are influenced by a mixture of land uses, contained one or more pesticides or degradates. 

4 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 1

Implications

Pesticides and degradates are likely to be present at detectable levels throughout 

most of the year in streams that have substantial agricultural or urban land use in 

their watersheds.

Streams are more vulnerable to pesticide contamination than ground water 

in most hydrologic settings, as indicated by much more frequent detections in 

stream water. 

The frequent detection of pesticides and degradates in shallow wells in agricultural and urban areas indicates that ground water may merit special attention in 

these land-use settings. Shallow ground water is used in some areas for drinking 

water—and can also move downward into deeper aquifers. Early attention to 

potential ground-water contamination is warranted because the movement of 

ground water is usually slow and contamination is difficult to reverse.

Pesticide occurrence in streams and ground water does not necessarily cause 

adverse effects on aquatic ecosystems or humans. The potential for effects can 

be assessed by comparing measured pesticide concentrations with water-quality 

benchmarks, which are based on the concentrations at which effects may occur. 

•

•

•

•

NAWQA studies included assessments of water quality in the most 

important agricultural and urban settings in each Study Unit, with studies 

of urban areas focused mostly on residential areas.

Overview of Findings and Implications 5
Concentrations greater than 

human-health benchmarks

Mixed land uses

Stream water

Shallow ground water

Stream water

Shallow ground water

Stream water

Shallow ground water

Stream water

Major aquifers

none

none

9.6%

1.2%

6.7%

4.8%

1.5%

1.0%

Agricultural areas

Undeveloped areas

Urban areas

Percentage of stream sites or wells with

one or more pesticides exceeding 

a human-health benchmark

0 25 50 75 100

Figure 1–2. Concentrations of pesticides and 

degradates measured in streams and ground water 

usually were lower than human-health benchmarks 

for the pesticide compounds analyzed by NAWQA. 

Many of the wells sampled for ground-water 

studies, but none of the stream sites sampled, were 

sources of domestic or public water supplies during 

the study period.

Potential Significance to Human 

Health—

Pesticides seldom occurred at concentrations 

greater than water-quality benchmarks for human 

health (Chapter 6) 

A screening-level assessment of the potential significance of pesticides to human health 

was based on comparing measured concentrations in streams and ground water with waterquality benchmarks for human health. These 

human-health benchmarks were derived from 

standards and guidelines developed by USEPA 

for drinking water. 

Although none of the NAWQA stream sites 

is located at actual drinking-water intakes, comparison of time-weighted annual mean concentrations to human-health benchmarks provides 

perspectives on (1) the likelihood that some 

current drinking-water intakes on streams may 

withdraw water with pesticide concentrations 

that exceed a benchmark, and (2) the potential 

long-term significance of pesticides to the quality 

of water that may be used as sources of drinking 

water in the future. 

 Annual mean concentrations of pesticides 

in streams studied by NAWQA were seldom 

greater than human-health benchmarks (fig. 1–2). 

No streams draining undeveloped land, and 

only one stream in a watershed with mixed land 

uses, had an annual mean concentration greater 

than a human-health benchmark. The annual 

mean concentrations of one or more pesticides 

exceeded a human-health benchmark in about 

10 percent of the 83 agricultural streams and in 

about 7 percent of the 30 urban streams sampled 

by NAWQA. The 2 urban streams where benchmarks were exceeded are in Texas (diazinon) and 

Hawaii (dieldrin). Agricultural streams located in 

the Corn Belt (Illinois, Indiana, Iowa, Nebraska, 

Ohio, and parts of adjoining states) and Mississippi River Valley accounted for most concentrations that exceeded benchmarks, all involving 

atrazine (5 sites), cyanazine (4 sites), or dieldrin 

(2 sites). If, as examined in Chapter 6, the atrazine human-health benchmark were changed to 

values from the updated atrazine risk assessment 

(USEPA, 2003a), then there would be 2 sites 

rather than 5 sites with exceedances (although 

NAWQA did not measure 2 of the 3 degradates 

required for that benchmark).

Of pesticides accounting for most exceedances, atrazine use remains high, use of 

cyanazine has been reduced sharply since the 

mid-1990s (with corresponding decreases in 

stream concentrations; Chapter 8), and dieldrin 

and aldrin uses were discontinued before the 

1992–2001 study period. Changes through the 

study period in the frequency of benchmark 

exceedances by atrazine and cyanazine were 

consistent with changes in agricultural use. As 

described in Chapter 6, the proportion of agricultural stream sites in the Corn Belt with atrazine 

concentrations that exeeded the human-health 

benchmark was greater for streams sampled during 1998–2000 than for streams sampled during 

either 1993–1994 or 1995–1997. In contrast, 

most sites where cyanazine exceeded its benchmark were sampled during 1993–1994, and no 

sites that were sampled during 1998–2000 had 

exceedances of the cyanazine benchmark.

For perspective on the relevance of NAWQA 

findings to drinking-water supplies, NAWQA 

land-use classifications for 1,679 public watersupply intakes that withdraw water from streams 

in the United States indicate that 55 percent of 

the intakes withdraw water from streams that 

drain watersheds with predominantly undeveloped land, 32 percent from streams with mixed 

land use, 12 percent from streams with agricultural land use, and 1 percent from streams with 

urban land use. Although the watershed land uses 

of NAWQA sites and water-supply intakes were 

6 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 1

classified in the same way, NAWQA sites tend 

to have more agricultural and urban land in their 

watersheds than do water-supply intakes in the 

same land-use categories. 

Human-health benchmarks were exceeded 

less often in ground water than in streams 

(fig. 1–2). One or more pesticides exceeded 

a benchmark in about 1 percent of the 2,356 

domestic wells and 364 public-supply wells sampled among studies in the three land-use settings 

and major aquifers. In contrast to the streams that 

were sampled, however, these wells are sources 

of drinking water—commonly without treatment 

in the case of domestic wells and with variable 

amounts and types of treatment for public-supply 

wells. Shallow ground water sampled in urban 

areas had the greatest proportion of wells with 

concentrations of one or more pesticides that 

were greater than a benchmark, including 1 of 9 

public-supply wells, 3 of 17 domestic wells, and 

37 of 835 observation wells, for a total of about 

5 percent. About 1 percent of wells sampled in 

agricultural areas (shallow ground water) and in 

major aquifers had concentrations greater than 

one or more benchmarks. Wells with a concentration greater than a benchmark were scattered 

among 36 of the 187 ground-water studies. All 

concentrations greater than a benchmark were 

accounted for by dieldrin (72 wells), dinoseb (4), 

atrazine (4), lindane (2), and diazinon (1). 

Implications

Concentrations of pesticides measured in streams 

draining undeveloped and mixed-land-use watersheds indicate that public water-supply intakes on 

streams in these land-use settings, which compose 

about 87 percent of all intakes on the Nation’s 

streams, are unlikely to withdraw water with concentrations that are greater than a human-health 

benchmark.

The likelihood of pesticide concentrations exceeding 

a human-health benchmark in streams is greatest 

for those streams draining agricultural or urban 

watersheds, which account for about 12 and 1 percent, respectively, of public water-supply intakes on 

streams (based on NAWQA land-use classification). 

Such streams may warrant a priority for enhanced 

monitoring.

The likelihood of pesticide concentrations exceeding 

a human-health benchmark in a public-supply well 

or domestic well is low on the basis of NAWQA 

results. About 1 percent of such wells sampled by 

NAWQA in all land-use settings had a pesticide 

concentration greater than a benchmark—most 

frequently dieldrin, which is no longer used. 

•

•

•

Characteristics and Limitations of the ScreeningLevel Assessment of Potential Effects

The NAWQA screening-level assessment provides an initial perspective 

on the potential importance of pesticides to water quality in a national context 

by comparing measured concentrations with water-quality benchmarks. The 

screening-level assessment is not a substitute for risk assessment, which 

includes many more factors, such as additional avenues of exposure. The 

screening-level results are primarily intended to identify and prioritize needs 

for further investigation and have the following characteristics and limitations.

Most benchmarks used in this report are estimates of no-effect levels, 

such that concentrations below the benchmarks are expected to have a 

low likelihood of adverse effects and concentrations above a benchmark 

have a greater likelihood of adverse effects, which generally increases 

with concentration. 

The presence of pesticides in streams or ground water at concentrations 

that exceed benchmarks does not indicate that adverse effects are certain to occur. Conversely, concentrations that are below benchmarks do 

not guarantee that adverse effects will not occur, but indicate that they 

are expected to be negligible (subject to limitations of measurements 

and benchmarks described below).

The potential for adverse effects of pesticides on humans, aquatic life, 

and wildlife can only be partially addressed by NAWQA studies because 

chemical analyses did not include all pesticides and degradates. In addition, some compounds analyzed by NAWQA do not have benchmarks.

Most benchmarks used in this report are based on toxicity tests of individual chemicals, whereas NAWQA results indicate that pesticides usually occur as mixtures. Comparisons to single-compound benchmarks 

may tend to underestimate the potential for adverse effects for some 

sites. 

Water-quality benchmarks for different pesticides and media are not 

always comparable because they have been derived by a number of 

different approaches, using a variety of types of toxicity values and test 

species.

For some benchmarks, there is substantial uncertainty in underlying 

estimates of no-effect levels, depending on the methods used to derive 

them and the quantity and types of toxicity information on which they are 

based. This is especially true of fish-tissue benchmarks for the protection of fish-eating wildlife, for which there is no consensus on nationalscale benchmarks or toxicity values.

Estimates of pesticide exposure derived from NAWQA concentration 

measurements are also uncertain—particularly estimates of short-term 

exposure of aquatic organisms to pesticides in stream water. Generally, 

short-term average concentrations in stream water, such as 4-day values, are underestimated from NAWQA data. 

•

•

•

•

•

•

•

Overview of Findings and Implications 7
High benchmark

Low benchmark

Concentrations greater than wildlife 

benchmarks

87%

88%

88%

 Fish tissue

 Fish tissue

 Fish tissue

 Fish tissue

51%

88%

Percentage of stream sites with 

one or more organochlorine pesticide 

compounds exceeding a wildlife benchmark

4.8%

25%

34%

17%

Concentrations greater than aquatic-life

benchmarks

Agricultural areas Agricultural areas

Urban areas

Undeveloped areas Undeveloped areas

Mixed land uses Mixed land uses

 Stream water 57%

 Bed sediment 31%

 Stream water

 Bed sediment

 Stream water

 Bed sediment

 Stream water

 Bed sediment

83%

70%

13%

8.4%

42%

36%

0 25 50 75 100 0 25 50 75 100

Percentage of stream sites with

one or more pesticide compounds

exceeding an aquatic-life benchmark

Urban areas

Figure 1–3. Pesticide concentrations measured in stream water and bed sediment frequently exceeded waterquality benchmarks for aquatic life. Concentrations of organochlorine pesticide compounds measured in whole-fish 

tissue were frequently greater than benchmarks for fish-eating wildlife, although the wide range of results for low and 

high benchmark values indicates relatively high uncertainty in the potential for effects, mainly because of uncertainty 

in the benchmark for total DDT. 

Potential Significance to Aquatic Life 

and Wildlife—

Concentrations of pesticides were frequently 

greater than water-quality benchmarks for aquatic 

life and fish-eating wildlife (Chapter 6) 

 A screening-level perspective on the potential significance of pesticides to aquatic life and 

fish-eating wildlife was obtained by comparing 

concentrations measured in streams—including those in water, bed sediment, and whole 

fish—with water-quality benchmarks derived 

from guidelines established by USEPA, toxicity 

values from USEPA pesticide risk assessments, 

or selected guidelines from other sources.

Water—NAWQA findings for streams 

indicate that pesticides detected in water, most 

of which were in use during the study period, 

frequently exceeded aquatic-life benchmarks 

(fig. 1–3). Of 186 stream sites sampled nationwide:

57 percent of 83 agricultural streams had 

concentrations of at least one pesticide that 

exceeded one or more aquatic-life benchmarks 

at least one time during the year (68 percent of 

sites sampled during 1993–1994, 43 percent 

•

during 1995–1997, and 50 percent during 

1998–2000).

83 percent of 30 urban streams had concentrations of at least one pesticide that exceeded 

one or more aquatic-life benchmarks at least 

one time during the year (90 percent of sites 

sampled during 1993–1994, 100 percent 

during 1995–1997, and 64 percent during 

1998–2000).

42 percent of 65 mixed-land-use streams had 

concentrations of at least one pesticide that 

exceeded one or more aquatic-life benchmarks 

at least one time during the year (38 percent of 

sites sampled during 1993–1994, 40 percent 

during 1995–1997, and 46 percent during 

1998–2000). 

Streams in which concentrations of one or 

more pesticides exceeded an aquatic-life benchmark for water were distributed throughout the 

Nation in agricultural, urban, and mixed-land-use 

settings. In urban streams, most concentrations 

greater than a benchmark involved the insecticides diazinon (73 percent of sites), chlorpyrifos (37 percent), and malathion (30 percent). 

A potential revision of the acute invertebrate 

benchmark for diazinon from 0.1 micrograms per 

•

•

8 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 1

liter (μg/L) to 0.4 μg/L, as discussed in Chapter 

6, would reduce the percentage of urban streams 

with exceedances by diazinon from 73 percent to 

40 percent. As described in Chapter 6, all three of 

these insecticides exceeded aquatic-life benchmarks least frequently at urban sites sampled 

near the end of the study period (1998–2000), 

compared with sites sampled during 1993–1997.

Agricultural and nonagricultural uses of diazinon 

and chlorpyrifos have been restricted to varying 

degrees since 2001, as discussed for diazinon in 

Chapter 8.

In agricultural streams, most concentrations 

greater than a benchmark involved chlorpyrifos (21 percent of sites), azinphos-methyl (19 

percent), atrazine (18 percent), p,p'-DDE (16 

percent), and alachlor (15 percent). Findings for 

agricultural streams in the Corn Belt indicate that 

alachlor exceedances declined through the study 

period, with none during 1998–2000; atrazine 

exceedances increased, with the most frequent for 

sites sampled during 1998–2000; and chlorpyrifos exceedances varied through the study period, 

but were most frequent during 1998–2000.

Generally, insecticides most commonly 

exceeded benchmarks that are based on acute or 

chronic effects on aquatic invertebrates, or those 

that are based on ambient water-quality criteria for aquatic life. Herbicides most commonly 

exceeded benchmarks that are based on acute or 

chronic effects on vascular or nonvascular plants. 

Because of the wide variability in the number, 

type, and degree of benchmark exceedances 

among sites and the complexity of translating 

exceedances of screening-level benchmarks into 

specific potential for effects, the screening-level 

results should be used as the starting point for 

further site-specific investigation. 

Bed Sediment—Concentrations of organochlorine pesticide compounds measured in bed 

sediment were greater than one or more aquaticlife benchmarks at 70 percent of urban stream 

sites, 31 percent of agricultural sites, 36 percent 

of sites with mixed land use, and 8 percent of 

undeveloped sites (fig. 1–3). The geographic 

distribution of sites where aquatic-life benchmarks for bed sediment were exceeded is similar 

to findings for water in many respects, including urban streams throughout the country, and 

many agricultural and mixed-land-use streams 

in the Southeast, East, and irrigated areas of the 

West. In urban streams, aquatic-life benchmarks 

were most frequently exceeded by individual 

compounds in the DDT group or total DDT (58 

percent of sites), total chlordane (57 percent), and 

dieldrin (26 percent). Compounds in the DDT 

group are derived from 2 parent pesticides, DDT 

and DDD, and include several degradates and 

by-products (DDD is also a degradate of DDT). 

Total DDT is the sum of the concentrations of six 

individual compounds. Total chlordane concentration is the sum of concentrations of the cis and 

trans isomers of chlordane and nonachlor, plus 

the chlordane degradate oxychlordane. In agricultural streams, aquatic-life benchmarks were 

exceeded most often by individual compounds in 

the DDT group or by total DDT (28 percent of 

sites) and by dieldrin (8 percent).

Fish Tissue—Comparisons of concentrations 

of organochlorine pesticide compounds measured 

in whole fish with benchmarks for fish-eating wildlife indicate a wide range of potential 

for effects, depending on the type of wildlife 

benchmark used (fig. 1–3). Because there is no 

consensus on tissue-based benchmark values for 

wildlife, measured concentrations were compared 

with both the high and low benchmark values 

from the range available for each compound. 

The high benchmark values for fish tissue were 

exceeded most frequently in streams in the populous Northeast; in high-use agricultural areas in 

the upper and lower Mississippi River Basin; 

in high-use irrigated agricultural areas, such 

as eastern Washington and the Central Valley 

of California; and in urban streams distributed 

throughout the country. In urban streams, low 

benchmarks were exceeded most often by total 

DDT (88 percent of sites), dieldrin (18 percent), 

and total chlordane (10 percent). In agricultural 

streams, low benchmarks were exceeded most 

often by total DDT (87 percent of sites), dieldrin 

(11 percent), and toxaphene (9 percent). 

Implications

The screening-level assessment indicates that the most widespread potential 

impact of pesticides on water quality is adverse effects on aquatic life and 

fish-eating wildlife, particularly in streams draining watersheds with substantial 

agricultural and urban areas.

Assessment and management of potential effects on aquatic life and wildlife are 

complicated by the combined presence of (1) currently used pesticides and their 

degradates, and (2) organochlorine pesticide compounds derived from pesticides 

that, for the most part, had their uses cancelled prior to 1990. 

The widespread potential for adverse effects shown by the screening-level 

assessment—and the uncertainty in this potential because of the preliminary 

nature of the assessment and the complexity of pesticide exposure—

indicate a continuing need to study the effects of pesticides on aquatic life and 

wildlife under the conditions of pesticide exposure that occur in the environment. 

•

•

•

Overview of Findings and Implications 9
Frequently Detected Pesticides and 

Relations to Use—

Pesticides detected most frequently were among 

those used most heavily during the study period or 

in the past (Chapter 4) 

The pesticides detected most frequently in 

streams and ground water were primarily those 

with the greatest use—either during the study 

period or in the past—and with the greatest 

mobility and (or) persistence in the hydrologic 

system (fig. 1–4). 

The pesticides detected most frequently in 

stream water included: (1) five agricultural herbicides that were among the most heavily used 

during the study period—atrazine (and its degradate deethylatrazine), metolachlor, cyanazine, 

alachlor, and acetochlor; (2) five herbicides 

extensively used for nonagricultural purposes, 

particularly in urban areas—simazine, prometon, 

tebuthiuron, 2,4-D, and diuron; and (3) three of 

the most extensively used insecticides during the 

study period—diazinon, chlorpyrifos, and carbaryl (fig. 1–4). Simazine, prometon, diuron, 2,4-D, 

diazinon, and carbaryl, which are commonly used 

to control weeds, insects, and other pests in urban 

areas, were frequently found at relatively high 

levels in urban streams throughout the Nation. 

The use of individual pesticides often changes 

over time, and may have increased or decreased 

during or since the end of the study period. For 

example, the uses of diazinon and chlorpyrifos 

have been substantially restricted since 2001, 

and analysis of recent data for diazinon shows 

that concentrations in some streams have now 

declined as well.

The pesticide compounds detected most 

frequently in fish and bed sediment were historically used organochlorine pesticides, along 

with their degradates and by-products (fig. 1–4). 

Most organochlorine pesticides were heavily 

used during the 1950s and 1960s, but had their 

agricultural uses cancelled during the 1970s and 

remaining urban uses cancelled by the late 1980s. 

Some organochlorine compounds, however, 

persist in soils, sediment, and biota. Several compounds in the DDT group, chlordane compounds, 

dieldrin (from use of both dieldrin and aldrin), 

and heptachlor epoxide (degradate of heptachlor), 

were found most frequently. Although quantitative information on urban pesticide use is limited, 

the relatively high concentrations found in fish 

and bed sediment from urban streams indicate 

that historical use of these pesticides in urban 

areas was probably intensive.

Compared with streams, ground-water 

detections were dominated by fewer compounds—mainly those with relatively high 

mobility and persistence, which allows them to 

move greater distances to and within the groundwater flow system (fig. 1–4). The most prevalent 

pesticides in both agricultural and urban areas 

were the herbicides atrazine (and deethylatrazine), metolachlor, prometon, and simazine. 

Implications 

The correlations of the pesticides found most frequently in streams and ground 

water with the amounts and characteristics of pesticides used can help managers to anticipate and prioritize the pesticides most likely to affect water quality in 

different land-use settings.

For pesticides that are still being applied, reducing their use is likely to be an 

effective way to reduce their concentrations in the hydrologic system, particularly 

for streams (other approaches may also be effective). 

For organochlorine pesticide compounds derived from past use, management 

practices that control the erosion of soil may help to reduce their transport to 

streams.

•

•

•

10 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 1

Figure 1–4. The pesticide compounds detected most frequently 

in streams and ground water in agricultural and urban areas were 

mainly those with the most extensive use—either during the study 

period or historically—and those with the greatest mobility and (or) 

persistence in the hydrologic system.

Heptachlor epoxide

Dieldrin

Oxychlordane

trans-Nonachlor

cis-Nonachlor

trans-Chlordane

cis-Chlordane

o,p'-DDT

o,p'-DDD

o,p'-DDE

p,p'-DDT

p,p'-DDD

p,p'-DDE

Total DDT

Total

chlordane

Agricultural

Atrazine

Deethylatrazine

Metolachlor

Cyanazine

Alachlor

Acetochlor

Simazine

Prometon

Tebuthiuron

2,4-D

Diuron

Diazinon

Chlorpyrifos

 Carbaryl

Atrazine

Deethylatrazine

Metolachlor

Cyanazine

Alachlor

Acetochlor

Simazine

Prometon

Tebuthiuron

2,4-D

Diuron

Diazinon

Chlorpyrifos

Carbaryl

Urban

Stream Water

Frequently detected pesticide 

compounds

Herbicides Insecticides Herbicides Insecticides

Organochlorine compounds

25 50 75 100 25 50 75 100

100 100

25 50 75 100 25 50 75 100

Ground Water

Percentage of time detected

Fish Tissue

Percentage of samples with detections

Percentage of samples with detections

1_04freq

25 50 75 25 50 75

Agricultural Urban

Agricultural Urban

The most intensive pesticide applications are in 

agricultural and urban areas, including substantial use for 

home, lawn, and garden pest control in residential areas 

(photograph ©2003 Corbis [top]).

Overview of Findings and Implications 11
Geographic Patterns—

Patterns of pesticide occurrence in streams 

primarily followed the distribution of use, whereas 

patterns in ground water were more affected by 

management practices and natural susceptibility 

to contamination (Chapter 4) 

The types and concentrations of pesticides 

found in agricultural streams primarily reflect 

the geographic distributions and intensity of 

use, along with additional influences by climate, 

soil characteristics, and water-management 

practices. For example, geographic patterns in 

stream concentrations of atrazine, metolachlor, 

simazine, acetochlor, 2,4-D, chlorpyrifos, and 

diazinon directly correlate with where they are 

used on crops. Some of the highest concentrations of atrazine were observed in streams within 

the Corn Belt and other areas where corn is a 

primary crop and where the herbicide is most 

heavily used (fig. 1–5). Total DDT was found at 

some of the highest concentrations in bed sediment and fish in parts of the Southeast, where 

DDT was historically used on cotton, tobacco, 

and peanuts, as well as in parts of California, 

Oregon, and Washington, where it was used 

extensively on orchards, potatoes, vegetables, and 

specialty crops. Dieldrin, on the other hand, was 

found most frequently and at some of the highest 

concentrations in the Corn Belt, where aldrin and 

dieldrin were extensively applied to corn.

The geographic distribution of pesticides in 

ground water also is influenced by the distributions of land use and pesticide use, but is more 

strongly affected by natural features, such as 

hydrogeology and soil characteristics, and by 

agricultural management practices, such as irrigation and drainage. For example, ground water is 

more susceptible to contamination in areas where 

the soil and unsaturated zone are more permeable 

than in areas where they are less permeable. A 

management practice that can influence pesticides in ground water is the use of subsurface 

tile-drain systems, which are buried networks 

of perforated pipes that collect shallow ground 

water for the purpose of lowering the water table 

and draining water-logged soils, as well as other 

subsurface drainage systems. These drain systems may reduce pesticide levels in underlying 

ground water by diverting shallow ground water 

to surface waters.

Detection frequencies of atrazine (fig. 1–6), 

metolachlor, and simazine generally were highest 

in ground water sampled in areas with permeable soils and geologic formations in parts of 

Atrazine use

Atrazine

concentrations

(in micrograms per liter, μg/L)

very low

low

medium

high

low ( 0.05)

medium (0.05 – 0.5)

high ( 0.5)

!

!

!

!

!

!

!

!! !! !!

!

!

! !!

!

! ! !!

! !

!

! !! ! !

! ! !! !

! ! ! !

!

!

!

!

! !!

! !

!

!

!

!!!

!!

! !

! !

!

!! ! !! !! !!!! !!

! ! !! ! !

!

!

Atrazine in agricultural streams

Figure 1–5. The concentrations of atrazine 

measured in agricultural streams correlated with 

the distribution of its use on crops—primarily corn. 

Some of the highest concentrations occurred in 

the corn-growing areas of Illinois, Indiana, Iowa, 

Nebraska, and Ohio. 

Atrazine use

Atrazine

detection frequency

very low

low

medium

high

not detected or 0.01 μg/L

low (10 percent)

medium (10 – 25 percent)

high ( 25 percent)

!

!

! ! !

! !

!

!

! !

!

! ! ! ! !

!!

!

! !

!

!

!! !!

! ! ! !

!

!

!

! ! ! !!! ! !

!! ! !

!

! ! !

!

!

Atrazine in shallow ground water 

in agricultural areas

Figure 1–6. Atrazine was detected most frequently 

in shallow ground water in agricultural areas where 

soils and the underlying unsaturated zone are highly 

permeable and use is moderate to high, such as 

in parts of Iowa, Minnesota, Pennsylvania, and 

Wisconsin.

12 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 1

the country where these compounds are used for 

corn production—such as parts of Iowa, Minnesota, Pennsylvania, and Wisconsin. In contrast, 

these herbicides were found less frequently and 

at lower concentrations in ground water within 

many areas sampled in the central Corn Belt, 

despite some of the highest use in the Nation. 

This apparent anomaly, which has also been 

noted by other studies, is probably caused by the 

relatively impermeable soils and glacial till that 

cover much of this region, combined with the 

resulting widespread use of subsurface drainage systems. As observed for streams, each 

pesticide has a unique pattern and story regarding its occurrence in ground water, in large part 

resulting from its use on particular crops and its 

characteristic mobility and persistence. Pesticide 

properties more strongly control the occurrence 

of pesticides in ground water than in streams, 

however, because longer travel times in ground 

water and prolonged contact with soil and aquifer 

materials reduce concentrations of pesticides or 

degradates with low persistence or mobility.

Implications 

Pesticide occurrence in streams can be largely anticipated from the geographic 

distribution of land use, crops, and associated chemical use. Other factors, such 

as soil and runoff characteristics, also influence the amount and timing of the 

transport of pesticides to streams, but these factors are generally less important 

than the amount used in determining pesticide concentrations in streams. 

Compared with streams, natural features and management practices are more 

important considerations for anticipating the occurrence of pesticides in ground 

water. Ground water is most susceptible to contamination in areas where soils 

and the underlying unsaturated zone are most permeable and drainage practices 

do not divert recharge to surface waters. 

The entire hydrologic system and its complexities need to be considered in evaluating the potential for pesticide contamination of streams and ground water. 

Some hydrologic settings where ground water is least vulnerable to contamination are those where streams are most vulnerable, and vice versa. For example, 

subsurface drains may help protect deep ground water, but increase pesticide 

transport to streams. 

•

•

•

Different pesticides are applied during different seasons 

in each region of the country. In the San Joaquin Valley, 

California, many orchards were sprayed with diazinon during 

the winter when they are dormant, whereas herbicides 

were applied to corn fields before and after spring planting 

throughout much of the Corn Belt (photograph by Dave Kim, 

California Department of Pesticide Regulation [left]). 

Overview of Findings and Implications 13
also may vary because of differences in the timing of local water-management practices, such as 

irrigation and reservoir releases.

Implications 

Effective management of streams may require 

increased monitoring—including high-frequency 

sampling during seasons when intense pesticide 

use coincides with periods of high runoff—so that 

the periods with the highest pesticide concentrations are adequately characterized.

Seasonal patterns in pesticide concentrations are 

important to consider in managing the quality of 

drinking water withdrawn from streams in agricultural and urban settings. Knowledge of seasonal 

patterns may help managers to adapt treatment 

strategies, or avoid or minimize withdrawals in 

favor of alternative sources of water, during highconcentration seasons.

Seasonal patterns may result in adverse effects on 

aquatic life in some streams. Both acute and chronic 

aquatic-life benchmarks for water were most 

frequently exceeded during seasonal periods of high 

concentrations. Concentration pulses of some pesticides during sensitive stages of aquatic life cycles 

may have the greatest effects in some streams, and 

site-specific assessments may be required. 

•

•

•

Figure 1–7. Atrazine concentrations in the White River, Indiana, were typical of streams throughout 

much of the Corn Belt, following similar seasonal patterns every year. Concentration pulses corresponded to the timing of runoff events and atrazine use on corn fields each spring. In contrast, 

seasonal patterns in concentrations of chlorpyrifos were more variable because of its more sporadic use.

0.001

1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

0.01

0.1

1

10

100

Pesticide concentration, in micrograms per liter

Atrazine

Chlorpyrifos

Seasonal patterns in the White River, Indiana

Streamflow

Seasonal Patterns—

Pesticide concentrations in streams followed 

distinct seasonal patterns (Chapter 5) 

Pesticide concentrations in stream water 

vary through the year, usually characterized by 

long periods with low or undetectable concentrations of most pesticides, punctuated by seasonal 

pulses of much higher concentrations. The timing 

and magnitude of seasonal pulses were correlated 

with the timing and intensity of pesticide applications, the frequency and magnitude of runoff 

from rainstorms or snowmelt, and the timing and 

distribution of land-management practices such 

as irrigation and artificial drainage. Concentrations in agricultural streams generally were 

highest during periods of runoff resulting from 

precipitation or irrigation that occurred soon 

after pesticide applications—a combination that 

causes seasonal patterns that are unique to each 

region. Spray drift and other modes of atmospheric transport can also be sources of pesticides to streams during high-use periods within 

an agricultural region. Most streams that drain 

farmland in the Corn Belt and other corn-growing 

areas, for example, had elevated concentrations 

of herbicides during spring runoff that followed 

applications (fig. 1–7). In contrast, agricultural 

streams in parts of the San Joaquin–Tulare Basin 

had high concentrations of diazinon during the 

winter, resulting from applications on dormant 

almond orchards followed by rainfall. Patterns 

14 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 1

Figure 1–8. Pesticides commonly occurred in 

streams and ground water as mixtures. For example, 

agricultural stream samples contained 10 or more 

different pesticides or degradates more than 20 

percent of the time.

Number of pesticide compounds in sample

1 2 3 4 5 6 7 8 9 10

0

25

50

75

100

Percentage of time (streams) or 

samples (ground water) with 

equal or greater number of detectable

pesticide compounds

Land use

Ground

water

Agricultural

Urban

Mixed

Streams

Frequency of mixtures in water

Mixtures—

Pesticides were most commonly detected as 

mixtures of multiple pesticide compounds 

(Chapter 5) 

Samples from streams in areas with substantial agricultural or urban land use almost always 

contained mixtures of multiple pesticides and 

degradates (fig. 1–8). More than 90 percent of the 

time, water from streams with agricultural, urban, 

or mixed-land-use watersheds had detections of 

2 or more pesticides or degradates, and about 20 

percent of the time they had detections of 10 or 

more. In addition, samples of fish tissue and bed 

sediment from most streams contained mixtures 

of historically used organochlorine pesticides and 

their degradates and by-products.

Mixtures were less common in ground 

water than in streams, which is consistent with 

the lower frequencies of detection for individual 

pesticide compounds. Nevertheless, 47 percent 

of shallow wells in agricultural areas and 37 

percent of shallow wells in urban areas contained 

2 or more detectable pesticides or degradates. 

Less than 1 percent had detections of 10 or more 

compounds. 

The environmental significance of mixtures 

is ultimately determined by the specific combinations of individual compounds—known as 

“unique mixtures”—their concentrations and 

combined toxicity, and how often and where they 

occur. A unique mixture is a specific combination 

of 2 or more compounds, regardless of the presence of other compounds. Thus, a single sample 

with several pesticides contains many unique 

mixtures. Depending on the specific compounds, 

the toxicity of a mixture to a particular type of 

organism may result from additive effects among 

the compounds, independent effects, antagonistic 

effects (less than additive), or synergistic effects 

(greater than additive). Each of these toxicity 

models, except for the antagonistic model, usually results in a toxicity of the mixture that is 

greater than any of its individual components.

More than 6,000 unique 5-compound mixtures were found at least 2 percent of the time in 

agricultural streams (only 1 unique 5-compound 

mixture was found in ground water). Evaluating 

the potential significance of mixtures can be simplified, however, because many mixtures do not 

occur very often at high concentrations, and the 

most frequently occurring mixtures are composed 

of relatively few pesticides. For example, the 

number of unique 5-compound mixtures found 

in agricultural streams is less than 100 when 

only concentrations greater than 0.1 micrograms 

per liter (μg/L) are considered. More than 30 

percent of all unique mixtures found in streams 

and ground water in agricultural and urban areas 

contained the herbicides atrazine (and deethylatrazine), metolachlor, simazine, and prometon. 

The insecticides diazinon, chlorpyrifos, carbaryl, 

and malathion were common in mixtures found 

in urban streams.

Implications 

Because of the widespread and common occurrence of pesticide mixtures, 

particularly in streams, the total combined toxicity of pesticides in water or other 

media often may be greater than that of any single pesticide compound that is 

present.

Continued systematic assessment is needed of the potential toxicity of pesticide 

mixtures to humans, aquatic life, and wildlife. NAWQA information on the occurrence and characteristics of mixtures can help to target and prioritize toxicity 

assessments. 

•

•

Overview of Findings and Implications 15
Degradates—

Concentrations of degradates were often 

greater than concentrations of parent pesticides 

(Chapter 5) 

Once released into the environment, pesticides undergo many types of transformation reactions that create degradates. Factors that govern 

the formation and distribution of degradates 

in the hydrologic system include the use and 

persistence of parent pesticides, the persistence 

and mobility of the degradates, and the physical, 

chemical, and biological conditions in the environment. In many cases, transformation results 

in the conversion of the parent compound to a 

compound that is less toxic, but some degradates 

have toxicities that are similar to, or greater than, 

that of their parent pesticide.

 Some degradates were found more frequently and at higher concentrations than their 

parent pesticide. For example, DDT, which 

was first used more than 50 years ago and was 

discontinued about 20 years before this study 

1996 1997 1998Application period

100

10

1

0.1

Concentration,

in micrograms per liter

Total parent

herbicides

Streamflow

Herbicides and their degradates in 

the Iowa River, Iowa

Total herbicide

degradates

Figure 1–9. The total concentration of degradates 

commonly exceeded the total concentration of parent 

herbicides (acetochlor, alachlor, atrazine, cyanazine, 

and metolachlor) in the Iowa River. 

began, was detected in fish from about 30 percent 

of agricultural streams sampled by NAWQA, 

whereas DDE, a more stable degradate of DDT, 

was detected in fish from 90 percent of sampled 

agricultural streams. Atrazine, the most heavily 

used herbicide in the Nation during the study 

period, was found together with one of its several 

degradates, deethyl­atrazine, in about 75 percent of stream samples and about 40 percent of 

ground-water samples collected in agricultural 

areas across the Nation. In the Eastern Iowa 

Basins, where NAWQA conducted special studies of herbicide degradates, an average of nearly 

85 percent of the total mass of herbicide compounds in stream samples was composed of 10 

degradates of the herbicides acetochlor, alachlor, 

atrazine, cyanazine, and metolachlor (fig. 1–9). 

The summed concentrations of degradates were 

more than 10 times higher than the summed 

concentrations of their parent compounds during 

much of the year, and the degradates accounted 

for the largest proportion of pesticide compounds 

that are transported by the Iowa River to the Mississippi River. 

Degradates are particularly important in 

ground water, which moves relatively slowly 

through soils and aquifers, providing the 

extended time and conditions favorable for 

transformation of pesticides to their degradates. 

Ground water in the Delmarva Peninsula, for 

example, contained degradates of alachlor and 

metolachlor at median concentrations 10 times 

higher than those of the parent herbicides. 

Degradates in ground water can ultimately reach 

streams when ground-water discharge contributes to streamflow. In the Iowa River, substantial 

transport of herbicide degradates occurred during 

low streamflow conditions (fig. 1–9). 

Implications

Pesticide degradates should continue to be considered and accounted for in 

assessments of pesticide exposure and in evaluating the potential effects of 

pesticides on humans, aquatic life, and wildlife.

Enhanced assessments of the occurrence and behavior of degradates in the 

hydrologic system require improved coverage of degradates in water-quality 

monitoring and continued research on pesticide transformations and transport 

in the hydrologic system. Enhanced assessments would supplement the toxicity 

testing of major degradates now required by USEPA as part of risk assessments 

for pesticide registration. 

•

•

16 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 1

Less than 5%

Probability of exceeding 3 μg/L atrazine

5% or greater

1_10

Prediction of atrazine in streams

Figure 1–10. Streams predicted to have a 5 percent 

or greater chance of having annual mean atrazine 

concentrations that are higher than its human-health 

benchmark of 3 μg/L—USEPA’s Maximum Contaminant Level 

for drinking water—are located throughout much of the 

Corn Belt and in other high-use areas such as the southern 

Mississippi River Valley. These estimates were based on 

1997 atrazine use for agriculture and will change if use 

changes.

Prediction—

Extensive data and improved understanding 

enable prediction of pesticide occurrence and 

concentrations for streams and ground water 

where they have not been measured (Chapter 7) 

NAWQA data from 1992 to 2001 are sufficiently extensive to support statistical models 

that can be used to estimate the concentrations 

or occurrence of some pesticides in streams 

and ground water where they have not yet been 

assessed. Such spatial extrapolation is fundamental to extending NAWQA’s targeted local 

and regional studies to a comprehensive national 

assessment. The statistical models were developed from measured pesticide concentrations, 

together with information on key factors and 

processes that affect pesticide occurrence, including pesticide use and land use, climate and soil 

characteristics, and other features. 

The NAWQA approach to extrapolation 

for streams is illustrated by a model used to 

estimate concentrations of atrazine in stream 

water, specifically the likelihood that the annual 

average atrazine concentration in any particular 

stream in the Nation would exceed a humanhealth benchmark of 3 μg/L (fig. 1–10). The 

human-health benchmark used for atrazine is the 

USEPA drinking-water standard, or Maximum 

Contaminant Level (MCL). Predictions are for 

annual mean concentrations in untreated stream 

water (including consideration of predictive 

uncertainty), regardless of whether a stream is 

presently a source of drinking water. Atrazine 

concentrations were predicted to be highest in the 

Corn Belt and parts of the southern Mississippi 

River Valley, where use is high and natural features favor the transport of pesticides by runoff 

to streams. In these areas, many streams are estimated to have more than a 5-percent chance of 

having a mean annual concentration of atrazine 

that is greater than the benchmark (shown in red 

in figure 1–10). In other words, more than 1 out 

of 20 of these streams are predicted to have mean 

concentrations greater than the human-health 

benchmark, and thus, may not be suitable as 

sources of drinking water without the use of strategies to lower concentrations. Similar analyses 

can be developed for other probability criteria or 

concentration estimates. 

Implications

The development of national-scale predictive models with quantified reliability 

is increasingly possible for some pesticides, particularly for streams. Expanding 

this capability is a critical step for national water-quality assessment, as well as 

for cost-effective management of water resources, because both require more 

information (compounds, places, and times) than can be directly measured under 

current technology and budget constraints.

Model estimates can be used to identify locations that have the greatest likelihood of water-quality problems and that are, therefore, the highest priority for 

additional monitoring.

Future success with development and application of statistical models—as well 

as more complex simulation models—will depend upon continued, carefully 

targeted monitoring of pesticide concentrations in the hydrologic system, coupled 

with continued and improved collection of supporting data on pesticide use, 

natural features, and other explanatory factors needed to update and validate the 

models. 

•

•

•

Overview of Findings and Implications 17
Figure 1–11. Total DDT concentrations in 

whole fish (wet weight) from rivers and streams 

throughout the Nation that drain watersheds 

with mixed land uses decreased rapidly from 

the 1960s through the 1970s, and then more 

slowly during the 1980s and 1990s.

500

0

1,000

1,500

2,000

2,500

3,000

3,500

4,000

1969 1974 1979 1984 1989 1994 1999

Total DDT concentration in fish, 

in micrograms per kilogram

Median

90th percentile

Declining DDT concentrations in fish

Trends—

A first look at trends shows examples of both 

decreasing and increasing levels of pesticides in 

streams and ground water (Chapter 8) 

NAWQA results from 1992 to 2001 provide 

a framework for assessing whether pesticide 

levels in the Nation’s streams and ground water 

are increasing or decreasing over time. For many 

pesticides and locations, it is too early to discern 

changes because historical data are insufficient 

to measure trends. Some trends, however, are 

already evident and others are just emerging.

The most complete story of trends in 

response to regulatory action and reduced 

pesticide use is the decline in concentrations of 

organochlorine pesticide compounds that followed reductions in use during the 1960s and 

bans on their uses in the 1970s and 1980s. For 

example, concentrations of total DDT in fish 

decreased rapidly from the 1960s through the 

1970s, then more slowly during the 1980s and 

1990s, as documented by data for 1969–1986 

from the U.S. Fish and Wildlife Service (Schmitt 

and Bunck, 1995), and by data for the 1990s 

from NAWQA (fig. 1–11). The trends in concentrations of organochlorine compounds in fish, 

however, also show that responses to reductions 

in sources can take a long time for chemicals that 

are persistent in the environment.

More recent regional changes in corn herbicide use have resulted in 

corresponding trends in concentrations in Corn Belt streams. For example, 

in response to the partial replacement of alachlor by the new herbicide acetochlor in 1994, streams quickly—generally within 1 to 2 years—showed 

increasing acetochlor concentrations and decreasing alachlor concentrations. Findings show that concentrations of relatively mobile and short-lived 

pesticides in stream water will respond rapidly to changes in use—much 

more quickly than the less mobile and more persistent organochlorine compounds in fish tissue. 

Ground water responds more slowly than streams to changes in pesticide use—taking years and even decades for changes in quality to occur. A 

persistent pesticide or degradate can remain in ground water long after its 

use has been discontinued because of the slow rates of ground-water flow 

and the resulting long residence time of water and contaminants in groundwater flow systems. This is evident from a number of studies in different 

parts of the country. For example, bromacil remained at detectable levels 

in ground water in parts of Florida for several years after it was no longer 

used. Similarly, dieldrin, which was no longer used during the study period, 

was still detectable at concentrations greater than its human-health benchmark in 72 wells sampled by NAWQA. 

Continued NAWQA studies and monitoring will build on the baseline 

assessment established during the 1990s to assess trends in basins across 

the Nation. Assessment of trends is a primary objective during the second 

decade of the NAWQA Program when study areas are systematically reassessed and an increasing number of stream and ground-water sampling 

sites will have had 10 years of monitoring. Equally important, the NAWQA 

studies will continue to link changes in pesticide occurrence and concentrations over time with those factors that control the timing of trends, such as 

changes in pesticide use, land management, and natural factors. 

Implications

Increases or decreases in pesticide use can result in rapid corresponding changes in pesticide concentrations in stream water—

generally within 1 to 2 years. In contrast, pesticide occurrence 

in ground water, and the occurrence of persistent compounds in 

aquatic organisms or sediment, may change slowly—sometimes 

taking decades to respond to changes in use.

Long-term and consistent monitoring of pesticides in streams and 

ground water is essential for distinguishing actual trends from 

short-term fluctuations and for accurately tracking changes. 

Assessment of trends in stream-water concentrations of most 

currently used pesticides requires consistent annual data, with 

a particular focus on critical seasons of high use and transport. 

Assessment of trends in more persistent pesticides, such as 

organochlorine compounds in fish tissue, can rely on samples collected several years apart.

Assessment of trends in concentrations of pesticides in ground 

water usually requires estimation of ground-water age and an 

understanding of the ground-water flow system because of the 

slow rate of ground-water flow and the uncertainty in flow paths. 

•

•

•

•

18 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
The NAWQA assessment provides the most comprehensive analysis to date of pesticides in streams and ground water at the national 

scale and serves as a foundation for improving water-resource 

assessment and management. Nevertheless, major gaps in critical 

information about pesticides still persist and continue to present challenges to scientists, managers, and policy makers. As present-day 

knowledge is brought to bear on decision making, there is a continuing need to improve the data and scientific understanding required 

for future decisions. Some of the most important steps needed to fill 

information gaps for pesticides are outlined below: 

Improve tracking of pesticide use. Existing data on pesticide 

use are sparse (infrequent, with coarse geographic coverage) 

for agricultural uses and virtually nonexistent for nonagricultural uses. Given the direct relations between pesticide use and 

occurrence, improvement in the extent, frequency, and quality of 

quantitative data on agricultural use—and development of comparable and reliable data sources for nonagricultural pesticide 

use—would have major benefits for assessment and management of pesticides in streams and ground water.

Add assessments of pesticides not yet studied. Many important 

pesticides have not yet been assessed in the Nation’s streams 

and ground water using a nationally consistent approach 

because of budget constraints and limitations of current analytical methods. These pesticides include most fungicides and 

fumigants in current use, as well as many new or increasingly 

used herbicides and insecticides, such as glyphosate and pyrethroid insecticides. Pesticides targeted for analysis need to be 

re-evaluated regularly as use changes over time and new pesticides are introduced. 

Improve assessment and understanding of degradates. Closely 

related to the gaps in pesticides that have been assessed, are 

the even greater gaps in information about the distribution of 

degradates in streams and ground water. Specific needs include 

the development of analytical capabilities for measurement of 

a broader suite of pesticide degradates, continued research on 

pesticide transformations and the implications for transport and 

persistence in the hydrologic system, and improved assessment 

of potential exposure to degradates and their potential to affect 

humans, aquatic life, and wildlife. This information is needed to 

supplement the information on the toxicity and environmental 

behavior of major degradates that is now developed as part of 

the pesticide registration process.

•

•

•

Evaluate toxicities of mixtures. Existing standards and guidelines for exposure to individual pesticides may not address all 

potential effects because actual exposure is most often to mixtures of multiple pesticides and degradates. Additional research 

is needed regarding the toxicities of mixtures to humans, aquatic 

life, and wildlife.

Evaluate the performance of management practices. Evaluations 

are needed for making direct links between management practices—such as irrigation methods, subsurface drains, integrated 

pest management, and retention of wetlands and buffers—and 

the concentrations and transport of pesticides in streams and 

ground water. Field-scale studies have shown that certain 

management practices can influence pesticides in streams and 

ground water, but the effectiveness of these practices has not 

been systematically assessed at regional and national scales. 

Improve methods for prediction. Successful assessment and 

management of the Nation’s water quality requires a commitment not only to monitoring pesticides and their degradates in 

streams and ground water, but also to the continued development of predictive tools, such as statistical and simulation models. NAWQA assessments demonstrate that models can play an 

important role in the assessment of water quality and provide 

a cost-effective approach for extrapolating measured waterquality conditions to unsampled areas. Predictive capabilities 

are critical because the expense of monitoring prevents direct 

assessment of pesticides for all of the places and times required. 

Models, however, are successful only if they are developed 

and verified on the basis of measured data. Thus, the integration of monitoring and modeling, which is heavily emphasized in 

NAWQA’s second decade of assessments, is critical to expanding and improving methods for prediction. 

Sustain and expand long-term monitoring for trends. Long-term, 

consistent data for assessing trends is essential for tracking 

water-quality response to changes in pesticide use and management practices, for providing early warning of unanticipated 

problems, and for updating and improving models. The second 

decade of NAWQA assessments will include long-term monitoring of a broad range of pesticides and degradates in water at 

a national network of selected sampling locations, but the geographic coverage and range of pesticides measured should be 

increased in cooperation with other agencies.

•

•

•

•

Priorities for Filling Information Gaps

Overview of Findings and Implications 19

Chapter 1
20 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
2

Pesticides are used to control weeds, insects, and other pests 

in agricultural areas, urban areas, and a variety of other landuse settings. Once released into the environment, pesticides 

and their degradates can move through the hydrologic system 

to streams and ground water, where they may affect humans, 

aquatic life, or wildlife if concentrations occur at toxic levels. 

Many factors affect the transport and concentrations of pesticides in streams and ground water, including the intensity and 

distribution of their use; natural factors, such as climate and 

soil characteristics; and the physical and chemical properties of 

the pesticide compounds themselves.

This chapter provides a 

basic overview of the use of 

pesticides and the factors that 

control their transport and fate 

in streams and ground water, 

thus serving as a foundation for 

understanding the occurrence 

and significance of pesticides in 

the hydrologic system—topics 

that are examined in subsequent 

chapters.

Pesticide Primer
Pesticides and Their Uses

A pesticide is any substance used to kill or 

control insects, weeds, fungi, rodents, bacteria, 

or other unwanted organisms. Pesticides provide 

a range of benefits, including increased food 

production and reduction of insect-borne disease, 

but their use also raises questions about possible 

adverse impacts on the environment, including 

potential effects on drinking-water sources and 

aquatic life.

 All pesticide products contain one or 

more active ingredients, which are referred to as 

pesticides in this report. Most pesticide products 

also contain adjuvants, which are usually referred 

to as inert ingredients on product labels. Active 

ingredients specifically target the pest organism, 

whereas adjuvants are used to increase the effectiveness of the active ingredient. Adjuvants were 

not extensively assessed as part of NAWQA studies and are not addressed in this report, although 

some may have toxicological importance. As of 

1997, about 900 pesticides were registered in the 

United States for use in more than 20,000 different pesticide products (Aspelin and Grube, 1999). 

New pesticides are introduced every year—for 

example, typically 10 to 20 new active ingredients were registered each year from 1967 to 1997 

(Aspelin and Grube, 1999).

 Conventional pesticides include four major 

groups: herbicides, insecticides, fungicides, and 

a mixed group of fumigants, nematicides and 

other pesticides. An average of almost 1 billion 

pounds (lb) of conventional pesticides was used 

each year in the United States during the 1992–

2001 study period (fig. 2–1). NAWQA studies 

focused primarily on herbicides and insecticides, 

as described further in Chapter 3. In addition 

to conventional pesticides, about 4 billion lb of 

other registered pesticides are used each year, 

including chlorine disinfectants, wood preservatives, and other specialty products. These other 

types of pesticides were not included in NAWQA 

assessments.

Pesticide names used in this report are the 

scientific names of the active ingredients, such 

as atrazine and diazinon—rather than product 

names—in order to minimize confusion among 

the vast array of products and their names. For 

example, a 1994 analysis identified a total of 

7,340 different product names associated with 

386 commonly used pesticides—an average of 

19 different product names per pesticide (Milne, 

1995). 

Pesticides are released into the environment 

primarily through their application to agricultural 

lands, such as croplands and orchards, and for 

nonagricultural pest control, such as on lawns 

and gardens, commercial areas, and rightsof-way. In 2001, agriculture accounted for 76 

percent of total national use, with the remaining 

24 percent being applied for a wide range of nonagricultural purposes (Kiely and others, 2004). 

The nature and extent of pesticide use for 

agriculture in the United States has continually changed over the past 40 years. Total use 

for agriculture steadily increased from 1964 to 

1980—from less than 400 million to more than 

800 million pounds per year (lb/yr)—and then 

varied between about 700 and 800 million lb/yr 

from 1980 to 2001 (fig. 2–1). From 1980 to 2001, 

the use of herbicides and fungicides decreased 

Figure 2–1. Total agricultural use of pesticides—

more than half of which was accounted for by 

herbicides—increased from 1964 until about 1980 

and then varied between about 700 and 800 million 

lb/yr until 2001. Nonagricultural use of pesticides, 

much of which occurs in urban settings, remained 

relatively steady from 1964 to 1979, gradually 

declined from 1979 to 1998, and then increased 

through 2001. (Data from Donaldson and others, 

2002; and Kiely and others, 2004.)

400

300

200

0

100

100

0

200

300

400

500

600

700

800

900

1965 1970 1975 1980 1985 1990 1995 2000

Year

Annual use, in millions of pounds

Agricultural use

Trends in pesticide use

Nonagricultural use

1965 1970 1975 1980 1985 1990 1995 2000

All pesticides

Herbicides

Fumigants, nematicides,

other miscellaneous pesticides

Insecticides

Fungicides

22 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 2

slightly, insecticide use decreased by about half, 

and the combined use of fumigants, nematicides, 

and other pesticides increased. 

By comparison, nonagricultural use of pesticides remained relatively constant from 1964 to 

1979, in the range of 250–300 million lb/yr, and 

then gradually decreased to about 190 million 

lb/yr by 1998. From 1998 to 2001, nonagricultural use increased, driven primarily by increases 

in the amounts of herbicides, insecticides, and 

fungicides applied for home and garden pest 

control (Kiely and others, 2004). 

Trends in the use of individual pesticides 

applied for agricultural or nonagricultural purposes commonly vary from these overall patterns, 

depending on factors such as market conditions, 

regulatory actions, and the introduction of new 

pesticides or approaches (for example, genetically engineered crops or organic agriculture). 

Selected trends are discussed in Chapter 8 as part 

of an evaluation of changes in pesticide occurrence in streams and ground water. 

Most agricultural use of conventional pesticides is accounted for by fewer than 100 active 

ingredients. In 1997, for example, 25 herbicides 

accounted for approximately 92 percent of total 

herbicide use, 25 insecticides accounted for 91 

percent of total insecticide use, 25 fungicides 

accounted for 99 percent of total fungicide use, 

and 20 pesticides accounted for 100 percent of 

total fumigant and nematicide use (Gianessi and 

Marcelli, 2000). The national use of a pesticide, however, does not necessarily indicate 

the importance of its use in a particular area. 

For example, some pesticides are used in small 

quantities nationally (ranking them outside the 

top pesticides by weight), but are used intensively or frequently in certain areas. Pesticide use 

within a particular agricultural area is determined 

by many factors, including the types of weeds, 

insects, and other pests of concern, the potency 

and application rates of specific pesticides, 

climate, regulatory limits, and cost. These factors 

are closely related to the types of crops and the 

extent of their production. For example, more 

than 50 percent of all agricultural pesticide use 

in the Nation (by weight) is for pest control on 

only three crops—corn, soybeans, and cotton. 

Applications to corn alone account for about 

30 percent of total pesticide use in the United 

States (fig. 2–2). Estimates of pesticide use for 

agricultural purposes are available for more than 

200 pesticides by crop and county for all States 

except Alaska and Hawaii (Thelin and Gianessi, 

2000; see http://ca.water.usgs.gov/pnsp/).

Pesticides are used for nonagricultural 

purposes in urban and suburban areas to control 

weeds, insects, and other pests around homes 

and gardens, in parks and golf courses, along 

roads and other rights-of-way, and in commercial 

and industrial areas. Pesticides are also used for 

nonagricultural purposes in undeveloped and 

agricultural areas to control weeds, insects and 

other pests along fence rows and roadsides, in 

parks, and around rural residences. Nationally, 

total nonagricultural use is estimated to exceed 

the amount applied to any single crop except 

for corn (fig. 2–2). Data on nonagricultural use 

of pesticides, however, are much more limited 

than those for agricultural use. The only current 

published national data on nonagricultural use 

of specific pesticides are for 10 of the pesticides 

most commonly applied in each of two nonagricultural market sectors: (1) home and garden 

use by homeowners, and (2) use for industrial, 

Figure 2–2. Agricultural use of pesticides in 2001 was 

about three times greater than nonagricultural use. However, 

nonagricultural uses, such as applications to control weeds 

and insects in urban and suburban areas, were second only 

to corn when compared with individual crops. (Pesticide 

use estimates for individual crops are for 1997 and are from 

Gianessi and Marcelli, 2000; use estimates for total agricultural 

and nonagricultural uses are for 2001 and are from Kiely and 

others, 2004.)

Corn

Pesticide applications

Soybeans

Cotton

Wheat

Pasture

Sorghum

Rice

Potatoes

Apples

Fallowland

Citrus

Selected crops

0 10 20 30 40 50 60 70 80 90 100

National use, as percentage of total 

Insecticides

Herbicides

Fungicides

Nonagricultural

use

Agricultural

use

Pesticide Primer 23
commercial, and government applications (Kiely 

and others, 2004).

In addition to total use, the intensity of 

pesticide applications—expressed as the amount 

applied per unit of land area (such as pounds 

per acre)—also is an important consideration in 

assessing possible effects on water quality. For 

example, land devoted to corn production constitutes only about 8 percent of the total national 

area of agricultural and urban land uses, and yet 

pesticide applications to corn account for 30 

percent of total national use. In contrast, pasture 

lands constitute about 50 percent of the total 

national area of agricultural and urban land, but 

account for only 4 percent of the total pesticide 

use. Collectively, urban areas across the Nation 

cover about the same area of land as corn (about 

8 percent of the total agricultural and urban land 

area) and account for much of the 24 percent 

of total pesticide use that is attributed to nonagricultural purposes (Kiely and others, 2004). 

Estimates from the early 1990s indicate that the 

intensities of applications of herbicides to turf 

grass at sod farms, golf courses, and residences—

and of applications of fungicides to turf at golf 

courses—are greater than the intensities of applications to most crops (Barbash and Resek, 1996). 

Available information indicates a relatively high 

intensity of nonagricultural use in urban areas. 

The specific pesticides applied in a particular area differ by land use, crop type, and targeted 

pests. For example, during the 1992–2001 study 

period, certain pesticides, such as 2,4-D, diuron, 

diazinon, and chlorpyrifos, were more intensively used in urban and suburban areas across 

the Nation than in most agricultural settings. The 

types of crops largely determine which pesticides are applied in agricultural areas, resulting 

in distinct geographic patterns of use (fig. 2–3). 

The use of each pesticide during 1997 was 

estimated by combining the 1997 state-level use 

data reported by Gianessi and Marcelli (2000), 

with county crop acreages from the 1997 Census 

of Agriculture (U.S. Department of Agriculture, 

1999), using methods described by Thelin and 

Gianessi (2000). Use intensity was mapped for 

agricultural land using land-cover data from the 

early 1990s (Vogelmann and others, 2001) as 

described by Nakagaki and Wolock (2005). For 

example, molinate is used only on rice, which is 

grown primarily in the lower Mississippi River 

Valley and in parts of Louisiana, Texas, and 

California. Diazinon, atrazine, and diuron also 

have their own unique patterns of application. 

Considering the combined use of all pesticides, 

the overall intensity of pesticide application for 

agriculture is greatest in the croplands of the 

Corn Belt, the Mississippi River Valley, Florida, 

the coastal plain of the Southeast and Mid-Atlantic states, and irrigated areas of the West. 

Although the amounts and intensities of 

pesticide applications largely define the sources 

of pesticides to the environment in a particular 

location, the potential of a pesticide to affect 

water quality is also influenced by its sources and 

pathways in the hydrologic system, its chemical 

and physical properties (which determine mobility and persistence), and its toxicity to humans, 

aquatic life, and wildlife. A basic background on 

the influences of sources, transport pathways, and 

pesticide properties on the behavior of pesticides 

in hydrologic systems—and some of the potential 

implications for water quality—is summarized 

below and these factors are further explored in 

Chapters 4 and 5. The potential effects of pesticides on humans, aquatic life, and wildlife are 

examined in Chapter 6.

24 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Pesticide Primer 25

Chapter 2

Diazinon

(fruits, nuts, and vegetables)

Agricultural use of selected pesticides

Atrazine

(corn and sorghum)

Diuron

(cotton, citrus, and alfalfa)

Estimated 1997 agricultural use intensity,

in pounds per square mile per year

Insecticides Herbicides

< 0.09 

0.09 – 4.5

> 4.5 – 45

> 45

< 0.09 

0.09 – 4.5

> 4.5 – 45

> 45

Molinate

(rice)

2_03maps

Figure 2–3. The geographic distribution of the use of each pesticide follows the distributions of crops and pests for which it is applied. 

Examples for the 1997 agricultural use of the insecticide diazinon and the herbicides atrazine, diuron, and molinate show four distinctly 

different geographic patterns of use. (Use intensity for 1997 was mapped for agricultural land using methods described on p. 24.)
Sources and Pathways in the 

Hydrologic System

Pesticides, like most other water contaminants, enter the hydrologic system from point 

sources, which are associated with specific points 

of release, and from nonpoint sources, which are 

diffuse and widely dispersed. Nonpoint sources 

are the dominant sources of pesticides found in 

streams and ground water (fig. 2–4). Nonpoint 

sources include runoff to streams from agricultural and urban land, seepage to ground water in 

areas where pesticides are used, and deposition 

of pesticides from the atmosphere. Potential point 

sources of pesticides include pesticide manufacturing plants, mixing-and-loading facilities, 

spills, waste­water recharge facilities (wells or 

basins), waste disposal sites, and sewage treatment plants. Once pesticides and their degradates 

(new compounds formed by the transformation of 

a pesticide by chemical or biological reactions) 

reach the atmosphere, streams, or ground water, 

they move through the hydrologic system with 

air, water, or particles, depending on the chemical 

and physical properties of the compounds. 

Atmosphere

The atmosphere is an important part of the 

hydrologic system in which pesticides can be 

transported substantial distances from where 

they are applied (for example, Goolsby and 

others, 1997; Cromwell and Thurman, 2000). 

In fact, atmospheric transport can be global 

and is thought to be responsible for the detection of long-lived organochlorine pesticides 

such as chlordane, DDT, and dieldrin in remote 

areas of Scandinavia and the Arctic—locations 

where these pesticides probably were never used 

(Majewski and Capel, 1995; Nowell and others, 

1999; Hermanson and others, 2005). 

Pesticides can be transported from the 

atmosphere to streams and ground water with 

precipitation or the deposition of particles. For 

example, studies in the Upper Tennessee River 

Basin showed that atrazine and metolachlor 

Figure 2–4. Pesticides are transported to streams and ground water primarily by runoff and recharge. Nonpoint sources of pesticides 

originating from areas where they were applied—rather than point sources such as wastewater discharges—are the most widespread 

causes of pesticide occurrence in streams and ground water. (Modified from Majewski and Capel, 1995.)

REGIONAL TRANSPORT

DRY

DEPOSITION

EVAPORATION

SPRAY DRIFT

WIND EROSION

RECHARGE GROUND-WATER RECHARGE

DISCHARGE

TO STREAMS

RUNOFF RUNOFF

ENTRY

THROUGH

RECHARGE WELLS

FROM

STREAMS

RUNOFF

The atmosphere is often overlooked as a source 

of pesticides, which return to earth with 

precipitation and dry deposition and can reach 

streams and ground water.

Pesticides are transported to 

the atmosphere during and 

after application.

Pesticides are transported 

to ground water mostly by

recharge resulting from 

rainfall or irrigation within 

agricultural and urban 

areas where they are used.

Other locally important 

sources include transport 

down leaky well casings 

and contaminated streams 

that lose water to ground 

water.

Streams are particularly 

vulnerable to pesticide 

contamination because runoff 

from agricultural and urban 

areas flows directly into streams 

along with both dissolved and 

particle-associated pesticides.

Ground water is an important 

source of pesticides to some 

streams during low-flow periods, 

when ground water can be a 

major portion of streamflow.

PRECIPITATION

26 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 2

were frequently detected in Clear Creek in the 

Obed National Wild and Scenic River watershed 

(fig. 2–5). Seasonal patterns of occurrence of 

these two compounds matched those in nearby 

agricultural streams, indicating an atmospheric 

source from agricultural areas (Hampson and 

others, 2000). Similarly, metolachlor, EPTC, and 

atrazine were detected in rainwater within the 

Minneapolis–St. Paul metropolitan area—where 

they had not been applied—also indicating 

that these pesticides had been transported from 

nearby agricultural areas (Andrews and others, 

1998; Capel and others, 1998). All three pesticides were found in rainwater at concentrations 

that were usually higher than those measured in 

the ground water of the Minneapolis–St. Paul 

study area; thus, atmospheric inputs alone could 

have accounted for most of their occurrence in 

the ground water. 

Streams

Pesticides are transported from land into 

streams primarily by runoff or drainage resulting 

from rainfall or irrigation. These event-generated 

inflows to streams can occur by surface runoff, 

shallow subsurface flow, or flow through drainage ditches and subsurface tile-drain systems. 

Some compounds, such as atrazine, readily dissolve in and move with water. Other compounds, 

such as chlorpyrifos, more strongly associate 

with soil particles and organic matter and are 

transported primarily with eroded soil, particularly during times of high runoff from precipitation or irrigation.

Transport to streams is controlled, in large 

part, by the timing of precipitation and associated 

runoff and drainage relative to pesticide applications (Leonard, 1990). For example, figure 2–6 

shows elevated concentrations of atrazine in the 

White River during runoff events that occurred 

soon after spring applications in May and June, 

but lower concentrations in response to runoff 

events of comparable or greater magnitude at 

other times of the year when use was lower, 

such as in April or August. The phenomenon of 

high herbicide concentrations in spring runoff 

has been extensively documented, especially in 

the Midwest (for example, Thurman and others, 

1991).

Pesticides also enter streams with inflowing 

ground water—which can be a continuous source 

of pesticides and degradates throughout the year 

in some areas (Squillace and others, 1993). For 

example, during baseflow conditions on the 

Figure 2–5. Atrazine and metolachlor were frequently 

detected in Clear Creek in the Obed National Wild and Scenic 

River watershed (Upper Tennessee River Basin). The seasonal 

patterns in concentrations followed those of applications in 

nearby agricultural areas, suggesting atmospheric transport 

(Hampson and others, 2000).

0

0.03

0.06

0.09

Concentration,

in micrograms per liter

1997

Atrazine and metolachlor in Clear Creek, 

Tennessee

1998

March

April

May

June

July

August

September

October

November

March

April

May

June

July

August

December

September

January

February

Atrazine

Metolachlor

2_05

Figure 2–6. Atrazine concentrations in the White River (White 

River Basin) rose rapidly and peaked during the first runoff 

events that occurred soon after spring applications of atrazine 

in May and June. High runoff and streamflow either before or 

substantially after the high-use period typically did not result in 

similarly high concentrations (Carter and others, 1995).

0 0

5,000

10,000

15,000

20,000

30,000

25,000

35,000

40,000

45,000

2

4

6

8

10

12

Concentration,

in micrograms per liter

Streamflow, in cubic feet per second

May

June

July

August

October

November

April

May

June

July

August

September

December

January

February

March

1991 1992

Streamflow

Atrazine

Atrazine in the White River, Indiana

Delmarva Peninsula, Shedlock and others (1999) 

measured pesticide concentrations in streams 

that were similar to those found in nearby wells. 

Baseflow conditions occur during periods of minimal precipitation, when streamflow is dominated 

by ground-water discharge. Similarly, in Waikele 

Stream, which drains a watershed with mixed 

land use on the Island of Oahu, Hawaii, concentrations of the herbicides bromacil, atrazine, and 

diuron were highest during baseflow conditions 

(Anthony and others, 2004). 

Once in a stream, a pesticide may transform, be taken up by aquatic organisms, attach 

Pesticide Primer 27
to suspended particles and be deposited in bed 

sediment, or volatilize to the atmosphere—all 

resulting in losses of the parent compound from 

stream water. Model calculations by Capel and 

others (2001) for 39 high-use pesticides indicated 

that transformations are the dominant cause of 

pesticide losses from streams, accounting for 

most of the predicted losses for 27 of the 39 compounds, compared with 4 pesticides lost mostly 

by volatilization, and 8 pesticides lost by multiple 

processes.

Ground Water 

Pesticides reach ground water primarily in 

water that infiltrates the soil and passes through 

the underlying unsaturated zone to the water 

table. As with streams, most pesticide transport 

to ground water is driven by rainfall or irrigation when one or both result in ground-water 

recharge. Ground-water transport is different 

from transport in streams because only dissolved 

forms of pesticides and their degradates move 

substantial distances with ground water. Particlebound compounds are largely retained by soil 

and aquifer materials. In addition, transport of 

pesticide compounds to and within ground water 

is much less predictable than transport in streams 

because the flow of ground water is considerably 

slower and more complex than the flow of stream 

water.

Pesticides and their degradates can move 

readily to ground water through mobile zones, 

such as cracks, worm holes, or permeable sediments, but a portion of pesticide compounds is 

retained in immobile zones within the subsurface 

where flow is minimal. Pesticide compounds 

that are retained in immobile zones can be 

released gradually to ground water by diffusion 

and subsequent leaching, sometimes long after 

application. Much as a soap-filled sponge must 

be repeatedly rinsed and wrung before all of the 

soap is removed, the soil acts as a reservoir from 

which pesticides and their degradates continue 

to leach after application. As a result, pesticides 

may be found in ground water much sooner 

than expected after application (because of rapid 

movement through mobile zones), as well as for 

extended periods afterward (because of gradual 

release from immobile zones). 

Pesticides generally are detected most 

frequently in ground water where the permeability of the soil and the hydraulic conductivity 

of the aquifer are highest, allowing relatively 

rapid transport. Deep aquifers usually have less 

pesticide contamination than shallow ground 

water because: (1) it takes a long time—decades 

or more, in many cases—for water to move from 

the land surface to deep ground water (resulting 

in long residence times for ground water and any 

solutes it may contain); (2) long travel distances 

increase the likelihood that pesticides will transform or attach to aquifer materials; (3) protective, 

low-permeability deposits (which inhibit flow 

and transport) may be present between the land 

surface and deep aquifers; and (4) the mixing 

of water from complex flow paths over long 

distances and time periods tends to result in a 

mixture of land-use influences on the quality of 

deep ground waters, commonly including contributions from areas of undeveloped land. 

Influence of Pesticide Properties on 

Environmental Behavior 

The occurrence and transport of pesticides 

in the environment are strongly influenced by 

the chemical and physical properties that affect 

their persistence and partitioning. Persistence 

refers to the tendency of a compound to remain 

in its original chemical form in the environment. 

Partitioning is the process by which pesticides 

become distributed among different environmental media, such as water, sediment, biota, and air, 

generally resulting in higher concentrations in 

some media than in others. 

Pesticides with high persistence remain in 

their original chemical form in the environment 

for long periods, whereas those with low persistence rapidly transform following their release. 

Transformations proceed at widely varying rates, 

depending on the structure of the compound and 

environmental conditions (Barbash, 2004). As a 

result, the persistence of a pesticide—which is 

commonly expressed in terms of a half-life for 

transformation (the amount of time that it takes 

for half of the compound to transform)—can vary 

from hours to decades. Some transformations of 

pesticides in the hydrologic system result in degradates whose chemical properties, toxicities, and 

ultimate fate are not well known, although much 

information on the properties of major degradates 

has begun to emerge in recent years as part of 

pesticide registration studies in the United States 

and Europe (Sinclair and Boxall, 2003). Generally, persistent pesticides or degradates may be 

transported for long distances or accumulate in 

soils, sediment, or biota. In some cases—as with 

several of the historically used organochlorine 

pesticides—both long-distance transport and 

accumulation have been observed. 

28 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 2

The tendency of a pesticide or degradate to partition into water, sediment, or other 

media determines where it is most likely to 

be detected in the environment, and how it is 

transported through the hydrologic system. Two 

of the parameters used most often to describe 

the partitioning of a compound among environmental media are (1) the Henry’s law constant 

(KH), which describes partitioning between air 

and water, and (2) the soil organic carbon-water 

partition coefficient (Koc), which describes partitioning between water and the organic matter 

in soil or sediment. Values of KH and Koc for the 

pesticides and degradates that were detected most 

frequently in NAWQA studies are provided in 

Appendix 2. These values were compiled from 

a variety of sources, including the U.S. Environmental Protection Agency (USEPA) and the U.S. 

Department of Agriculture (USDA). A pesticide 

with a high KH is volatile and thus, primarily 

tends to reside in and be transported by air. As 

a result, such compounds are rarely retained for 

long in streams or soil, but if they reach ground 

water, these compounds may remain for substantial periods of time because there is comparatively little exposure to the atmosphere. A 

pesticide with a high Koc has a greater tendency 

to accumulate in soil or sediment than to remain 

dissolved in water. Because they associate more 

strongly with organic matter than with water, 

pesticides with high Koc values are sometimes 

referred to as hydrophobic. Compounds with low 

Koc values (which therefore tend to favor water 

over organic matter) are described as hydrophilic. 

As a result of their affinity for organic matter, the 

more persistent hydrophobic pesticides are likely 

to accumulate not only in soils and sediments, 

but also in fish, birds, mammals, and other biota 

(Nowell and others, 1999).

Understanding the factors that affect the 

persistence of a pesticide and its occurrence in 

different environmental media is key to evaluating and anticipating its potential effects on water 

quality. For example, atrazine, which moves 

readily with water and is relatively persistent, 

reached streams in 10 times greater proportions 

than EPTC, which is less persistent, more volatile, and, unlike atrazine, usually incorporated 

into the soil when applied. Nationally, an average 

of about 1 percent of the atrazine applied to the 

land in watersheds of sampled streams reached 

its associated stream outlet, as opposed to only 

about 0.1 percent for EPTC (fig. 2–7) (Larson 

and others, 1999; Capel and others, 2001). 

Similarly, NAWQA data from agricultural areas 

across the Nation indicate that pesticides with 

Atrazine

EPTC

Annual load in streams, in kilograms

100 1,000 10,000 100,000 1,000,000 10,000,000

100 1,000 10,000

1%

0.1%

0.01%

10%

1%

0.1%

0.01%

10%

100,000 1,000,000 10,000,000

1,000,000

10,000

100,000

100

1,000

10

1

0.001

0.01

0.1

1,000,000

10,000

100,000

100

1,000

10

1

0.001

0.01

0.1

Annual agricultural use, in kilograms

Transport of atrazine and EPTC in streams

Figure 2–7. The annual transport of atrazine in

streams was typically equivalent to about 1 percent

of the amount applied in a particular watershed

(percentages are shown by diagonal lines), whereas

annual transport of EPTC in streams, which is less

persistent and more volatile than atrazine, was less

than 0.1 percent of annual use for more than half of

the sites (Larson and others, 1999).

Pesticides with high KOC values attach to sediment particles 

and are transported during runoff events when suspendedsediment concentrations are high.

Pesticide Primer 29
Figure 2–9. In the Yakima River Basin, transport of 

pesticides with high Koc values (high affinity for soil 

particles) was greater in streams draining larger 

proportions of farmland that was rill irrigated, a 

practice that results in soil erosion. In contrast, the 

amount of rill irrigation had little effect on low-Koc

pesticides, which have lower affinities for soil 

particles and are not as affected by soil erosion. 

(modified from Fuhrer and others, 2004.)

Median pesticide transport, as a

percentage of amount applied

Percentage of agricultural land

 that was rill irrigated

10

1

0.1

0.01

0.001

0–4 5–32 33–73 74–100

Low-Koc pesticides (Koc 300)

High-Koc pesticides (Koc 300)

Rill irrigation and pesticide transport 

in streams

greater persistence in soil are more likely to be 

detected in shallow ground water than compounds that are less persistent (Barbash and others, 1999). For example, within areas of relatively 

equal use, atrazine (soil half-life of 146 days) 

was detected in shallow ground water much more 

frequently than the less persistent metolachlor 

(soil half-life of 26 days) (fig. 2–8). 

The effects of chemical and physical properties on the environmental behavior of pesticides 

were also illustrated by the differences between 

hydrophobic and hydrophilic pesticides in their 

transport within streams of the Yakima River 

100

0

20

40

60

80

100

20

0

40

60

80

0.001 0.01 0.1 1 10 100

0.001 0.01 0.1 1 10 100

Annual agricultural use intensity, 

in kilograms per square kilometer

3 studies

2 studies

Frequency of detection at or above 0.01 micrograms per liter, in percent 

Metolachlor (half-life = 26 days)

Atrazine (half-life = 146 days)

Atrazine and metolachlor in ground water

Figure 2–8. Frequencies of detection in 

shallow ground water beneath agricultural 

areas were usually greater for atrazine than for 

metolachlor given equivalent annual use rates, 

in part because of the greater persistence of 

atrazine in soil (Barbash and others, 1999).

Basin. Hydrophobic pesticides (Koc ≥ 300 milliliters per gram, mL/g) were found to be most 

readily transported in basins dominated by rill 

irrigation—a management practice that involves 

distributing large volumes of water across the 

land surface, and which often causes soil erosion 

(fig. 2–9). In contrast, annual loads of hydrophilic pesticides (Koc < 300 mL/g) were not 

affected by the extent of rill irrigation because 

the compounds are, for the most part, transported 

in the water rather than in suspended sediment 

(Fuhrer and others, 2004). 

30 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
NAWQA’s Approach to Pesticide Assessment 31

Chapter 3

NAWQA’s assessment of pesticides during 1992–2001 provides the most comprehensive analysis to date of pesticide 

occurrence in streams and ground water of the United States. 

The assessment followed a national study design in which the 

most important hydrologic systems were studied on a rotational 

schedule using nationally consistent sampling and analytical 

methods. This approach yields an understanding of waterquality conditions in a national context, while also supporting 

comparisons and assessments within and among individual 

watersheds, aquifers, and geographic regions. By systematically relating pesticide occurrence and transport to key factors 

that control contamination—such as pesticide use and properties, land use, hydrology, and other environmental features—

NAWQA provides information needed for managing pesticides 

within the Nation’s diverse land-use and environmental settings.

This chapter summarizes the 

primary features of the NAWQA 

study design, providing the 

context for understanding 

findings about pesticides in the 

Nation’s streams and ground 

water.

3

NAWQA’s Approach to Pesticide Assessment
Targeted Sampling Across the 

Nation’s Diverse Land Uses and 

Environmental Settings

This report is based on results from 

NAWQA’s first decade of water-quality assessments, which were completed on a rotational 

schedule from 1992 to 2001 in 51 major hydrologic systems across the country—referred to as 

Study Units—as well as the High Plains Regional 

Ground Water Study, using a nationally consistent study design (see p. iv–v). Assessments were 

conducted in 20 Study Units during 1992–1995; 

in 16 during 1996–1998; and in 15 Study Units 

during 1998–2001. Collectively, the 51 NAWQA 

Study Units and the High Plains Study cover 

a substantial portion of the Nation’s land area; 

account for more than 70 percent of total water 

use and more than 50 percent of the population 

served by public water supplies and domestic 

wells; and are representative of the Nation’s 

diverse landscapes, hydrologic systems, ecological resources, and land uses. 

The primary objectives of the NAWQA 

pesticide assessment were to determine: (1) the 

occurrence and concentrations of pesticides in 

streams (ranging from small streams to large 

rivers) and ground water; (2) where and when 

pesticides occur in relation to factors that govern 

their sources and transport in the hydrologic 

system; (3) whether any pesticides may be present at concentrations that could affect human 

health, aquatic life, or fish-eating wildlife; and 

(4) how concentrations are changing over time. 

To address these goals, NAWQA employed a 

targeted assessment focusing on studies of:

streams and shallow ground water in specific, 

relatively homogeneous land-use and environmental settings to relate pesticide occurrence 

to individual types of nonpoint sources; and

streams and major aquifers (regionally extensive aquifers that are important ground-water 

resources for water supply) in areas of mixed 

land uses to evaluate the integrated effects of 

multiple sources of pesticides on their occurrence and concentrations.

Details on the sampling design and analytical 

methods, as well as all data used in this report 

are available at: http://ca.water.wr.usgs.gov/pnsp/

pubs/circ1291/.

For the targeted assessment by land use, 

streams and shallow ground water were sampled 

in agricultural and urban areas, and in undeveloped areas dominated by forest or rangeland. 

As described in more detail below and in the 

accompanying sidebar, streams and ground water 

were sampled most intensively in agricultural 

and urban areas because of the importance of 

assessing pesticide occurrence in areas where the 

compounds are used most intensively. The agricultural areas are diverse in climate, geography, 

and crop types, and span coastal, desert, and temperate environmental settings. They include, for 

example, areas dominated by production of corn 

and soybeans in the Midwest; wheat and other 

grains in the Great Plains; mixed row crops and 

poultry in the East; rangeland in the Southwest; 

rice in Louisiana; pineapple in Hawaii; and areas 

of grain, fruits and nuts, vegetables, and specialty 

crops in California and the Pacific Northwest. 

The areas sampled in urban settings were primarily residential, typically with low-to-medium 

population densities (300 to 5,600 people per 

square mile). Some commercial or industrial 

areas also were included, but point sources and 

extensive industrial and downtown urban areas 

generally were not assessed. 

NAWQA Stream Assessment in a 

National Context

Potential land-use influences on the quality 

of water sampled at NAWQA stream sites were 

characterized by determining the proportions of 

each major land use within each stream’s contributing watershed. Table 3–1 lists the criteria used 

by NAWQA to classify each stream sampling site 

by its predominant land-use category. Streams 

classified as “mixed land use” drain mixtures of 

•

•

Table 3–1. Each stream sampled by NAWQA was classified according to 

the dominant land uses in its watershed. The land-use data set used for these 

classifications was an enhanced version of the USGS 1992 National Land 

Cover Data (NLCD), which classified land use for each 30-by-30-meter area of 

land in the conterminous United States. The original and enhanced versions 

of the NLCD are described, respectively, by Vogelmann and others (2001) and 

Nakagaki and Wolock (2005). 

Land-use classification Watershed land-use criteria

Agricultural > 50 percent agricultural land and ≤ 5 percent 

urban land

Urban > 25 percent urban land and ≤ 25 percent 

agricultural land

Undeveloped ≤ 5 percent urban land and ≤ 25 percent 

agricultural land

Mixed All other combinations of urban, agricultural, and 

undeveloped land

32 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Water-quality assessments by NAWQA, which is a single program 

among many local, State, and Federal programs, were not designed 

to address all of the Nation’s water-resource information needs and 

issues. Listed below are several characteristics and limitations of the 

NAWQA approach that are important to consider when interpreting 

the findings on pesticides presented in this report.

NAWQA assessments characterized the quality of the available, 

untreated water resources, and not the quality of drinking water 

(as would be done by monitoring water from water-treatment 

plants or from household taps). By focusing on the quality of 

streams and ground water in their present condition (ambient 

water quality), NAWQA complements many Federal, State, and 

local drinking-water monitoring programs.

NAWQA assessments did not focus on specific sites with known 

water-quality problems or narrowly defined “issues of the day,” 

but rather on the condition of the total resource, including 

streams and ground water in a wide range of hydrologic and 

land-use settings across the country. 

NAWQA assessments of pesticides focused primarily on nonpoint sources resulting from applications for pest management in 

agricultural, urban, and other land-use settings, although some 

sites—particularly those downstream from major metropolitan 

areas—also may be influenced by point sources, such as discharges from wastewater treatment plants.

NAWQA assessments targeted specific land-use settings that 

are most extensive or important to water quality in a wide range 

of hydrologic and environmental settings across the Nation. 

This targeted approach gives priority to understanding the most 

critical factors influencing water quality. Extension of results to 

national analysis, however, requires careful definition of each 

type of water resource and environmental setting for which conclusions are drawn and the use of statistical models to extrapolate results to resources that have not been measured.

•

•

•

•

USGS analytical methods were designed to measure concentrations as low as economically and technically feasible. Studies 

of contaminant occurrence and behavior benefit from the most 

information possible at all concentration levels, and such data 

help to identify emerging issues and to track changes in concentrations over time. By this approach, however, pesticides were 

commonly detected at concentrations far below Federal or State 

standards and guidelines for protecting water quality. Detections 

of pesticides do not necessarily indicate that there are appreciable risks to human health, aquatic life, or wildlife. The potential 

for such risks must be assessed by comparing measured concentrations with those that may cause adverse effects.

USGS methods for analyzing pesticides in water measured 

concentrations in filtered water samples and, thus, may underestimate concentrations of compounds that have strong affinities 

for suspended particles. The potential for underestimation is 

greater for stream water compared with ground water because 

of the generally greater amounts of suspended particles present 

in stream water—which are removed by filtration along with any 

pesticides contained in or on the particles.

Pesticide compounds analyzed in water by NAWQA included 

many of the most heavily used herbicides and insecticides, but 

they included only a fraction of all pesticides currently in use and 

few of their degradates. NAWQA findings provide insights about 

what to expect for pesticides and degradates that were not measured, but must be considered as only a partial assessment of 

currently used pesticides.

Organochlorine pesticide compounds analyzed by NAWQA in 

bed sediment and fish tissue are predominantly related to pesticides that were no longer in use by 1990. Of the pesticide compounds measured in bed sediment and fish tissue, only dacthal, 

endosulfan, lindane, methoxychlor, and permethrin were used 

during all or part of the study period.

•

•

•

•

Unique Features of the NAWQA Approach

NAWQA’S Approach to Pesticide Assessment 33

Chapter 3
two or more land-use settings and do not meet 

the criteria described in table 3–1 for individual 

agricultural, urban, or undeveloped settings. 

Land-use classifications were adjusted for a small 

number of streams that have watersheds with 

substantial areas that did not contribute streamflow during the study period. Most streams that 

were classified as agricultural, urban, or undeveloped also commonly have small amounts of other 

land uses in their watersheds. For example, and 

of particular importance to findings for pesticides, many streams classified as undeveloped 

have some agricultural or urban activity in their 

watersheds. 

Consistent with the sampling design, which 

was targeted by land use, the NAWQA pesticide 

findings discussed in this report generally are 

presented by land-use category. Aggregation of 

NAWQA findings for streams across all landuse categories would not accurately represent 

all streams in the conterminous United States 

(fig. 3–1), which were characterized by classifying the watersheds of all stream segments in the 

USEPA river reach file (Nolan and others, 2003) 

using NAWQA land-use criteria (table 3–1). 

For example, nearly 40 percent of the streams 

sampled by NAWQA were agricultural streams, 

whereas agricultural streams represent about 15 

percent of all streams in the conterminous United 

States. Furthermore, as shown in figure 3–1, agricultural streams represent only about 10 percent 

of all streams with public water-supply intakes. 

There are 1,679 public water-supply intakes on 

streams across the Nation for which land use 

could be characterized. The NAWQA design also 

over-represented urban streams and under-represented undeveloped streams compared with the 

national occurrence of streams in these land-use 

settings (fig. 3–1). 

Even when grouped by land-use category, 

the watersheds of NAWQA sampling sites that 

were classified as agricultural and urban still 

tend to have higher proportions of agricultural 

and urban land than most streams nationwide 

in the same land-use groups, as well as streams 

with public water-supply intakes in the same 

land-use groups. For example, about 25 percent 

of agricultural streams sampled by NAWQA had 

watersheds with more than 90 percent agricultural land, compared with about 18 percent of 

all agricultural streams in the United States with 

more than 90 percent agricultural land. This 

indicates that NAWQA estimates of pesticide 

occurrence for agricultural and urban streams (for 

water and, to a lesser degree, bed sediment) are 

likely, as groups, to be high relative to those for 

other streams nationwide in these same land-use 

classes. 

NAWQA Ground-Water Assessment in 

a National Context

NAWQA assessed pesticides in groundwater within specific land-use settings and in 

major aquifers with influences of a mixture of 

land uses. Land-use studies focused on shallow 

ground water primarily within agricultural and 

urban land-use settings, and to a lesser extent in 

undeveloped areas. Each of these studies involved 

the sampling of about 20 to 30 randomly located 

wells (using either existing or newly installed 

wells) within each targeted land-use area. Most 

of the wells selected or installed for the land-use 

studies sampled ground water from less than 

20 feet below the water table, thus indicating as 

directly as possible the influence of each land use 

on shallow ground-water quality.

Major aquifer studies involved the sampling 

of about 20 to 30 domestic or public-supply wells 

that withdraw water from aquifers or aquifer 

systems that are major current or future sources 

of water supply. Wells that were sampled for 

these studies were randomly selected throughout 

the areas underlain by each major aquifer, without regard to land use. Thus, the ground water 

sampled for the major aquifer studies reflects 

the effects of a mixture of different land uses 

and ground-water ages on water quality, often 

including water that recharged long distances 

Figure 3–1. The NAWQA design for stream assessments placed 

greater emphasis on sampling streams that drain agricultural and 

urban watersheds (as defined in table 3–1), relative to those in 

undeveloped watersheds. Streams sampled by NAWQA included 

higher proportions of agricultural and urban streams—and lower 

proportions of undeveloped streams—compared with all streams 

in the conterminous United States and those with drinking-water 

intakes. 

0

10

20

30

40

50

60

70

NAWQA water

sampling sites

NAWQA

bed-sediment

sampling sites

U.S.

streams

Agricultural

Undeveloped

Land use

Urban

Mixed

Drinking-water

intakes

on streams

Proportion of watersheds, 

in percent

Land-use characteristics of streams

34 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 3

0

10

20

30

40

50

60

70

NAWQA sampled wells All U.S. land area

Proportion of sampled wells

or land area, in percent

Agricultural

Undeveloped

Land use

Urban

Mixed

Land-use characteristics of NAWQA ground-water

studies compared with National land use

Figure 3–2. NAWQA’s targeted sampling design for ground water 

emphasized areas with urban, mixed, and agricultural land use—

and under-represented undeveloped areas—when compared with 

the national distribution of land uses. All wells sampled for major 

aquifer studies were classified as mixed land use for this graph.

from the sampled wells and in a variety of different land-use settings. The nature and extent of 

each major aquifer sampled for these studies is 

described in the summary report for the NAWQA 

Study Unit in which it occurs (see http://water.

usgs.gov/nawqa/). 

NAWQA findings for ground water, as for 

streams, are grouped by land use in this report. 

NAWQA’s targeted sampling design for ground 

water over-represented areas with urban and 

mixed land use, somewhat over-represented agricultural areas, and under-represented undeveloped areas when compared with the national distribution of these land-use settings by land area 

(fig. 3–2). Comparisons of land-use distributions 

for the NAWQA ground-water studies with those 

for the entire Nation were based on NAWQA 

land-use classifications (as defined in table 3–1) 

for every square kilometer in the conterminous 

United States. Although NAWQA agricultural 

and urban stream sites, as discussed above, tend 

to have greater proportions of agricultural and 

urban land in their watersheds than other streams 

in the same land-use groups, ground-water 

studies are not expected to have this tendency 

within these land-use groups because each well 

was selected by site-specific land-use criteria 

designed to meet the land-use objective for each 

study. Such site-specific control of land-use characteristics was not possible for streams because 

of the relatively large areas included in each 

watershed. 

Large streams and rivers required sampling 

from bridges, boats, or cableways.

Most wells sampled for major aquifer studies were 

existing water-supply wells.

NAWQA’s Approach to Pesticide Assessment 35
Sampling Design

The NAWQA national assessment of pesticides is based on results from the analysis of 

more than 10,000 samples of water, bed sediment, and fish tissue from thousands of locations 

within the 51 NAWQA Study Units. Water-soluble pesticides, most of which were in use during 

the study period, were assessed in stream water 

and ground water. Organochlorine pesticides, 

which are no longer used in the United States, 

but remain persistent in the environment, were 

assessed in bed sediment and fish tissue—environmental media in which they accumulate.

Stream Water

Water samples were collected at 186 stream 

sites for analysis of pesticides and degradates dissolved in water (fig. 3–3). The samples were collected from streams throughout the year, including high-flow and low-flow conditions. Sampling 

was most intensive during the time of highest 

pesticide use and runoff—generally weekly or 

twice monthly for a 4- to 9-month period. Most 

analyses in this report are based on 1 year of data 

for each site (generally representing the single 

most complete year of sampling) to give equal 

influence to each stream. Because of the rotational assessment approach, the most complete 

year of sampling for each stream ranged from 

1993 to 2000, depending on the particular site. 

Bed Sediment and Fish

Samples of bed sediment were collected 

at 1,052 sites (fig. 3–3) and fish-tissue samples 

were collected at 607 of the bed-sediment sites 

(plus 93 additional sites not shown on the map) 

for analysis of organochlorine pesticides and 

selected degradates and by-products. At each site, 

fine-grained surficial bed sediment (sieved to < 2 

millimeters [mm]) was collected from multiple 

depositional areas within a stream reach on a 

single date—usually during low-flow conditions—and combined into a single composite 

sample for chemical analyses. For fish, multiple 

individuals of the same species were collected at 

a site, also on a single date, and whole fish were 

composited for chemical analyses of tissue.

Ground Water

Water samples were collected from 5,047 

wells in 187 land-use and major aquifer studies 

for analysis of pesticides and degradates dissolved in water (fig. 3–3). For the land-use studies, most of the wells sampled were new or existing observation wells or domestic supply wells. 

The major aquifer studies focused almost exclusively on existing wells used either for domestic 

or public supply. Repeated sampling, such as that 

conducted at stream sites, was not included for 

ground water because of the comparatively slow 

rate of change in most ground-water systems, 

relative to streams. Data analyses were based on 

one sample per well.

Most wells sampled for agricultural and urban land-use 

studies were observation wells installed by NAWQA.

36 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 3

Figure 3–3. Stream sampling sites and ground-water studies in agricultural, urban, and undeveloped areas were 

distributed across the Nation’s diverse environmental settings to evaluate the occurrence of pesticides within areas 

of specific land uses. Pesticides also were assessed in streams and major aquifers that represent the water-quality 

effects of mixed land uses and varied environmental settings.

Ground-Water Studies

Agricultural (53 studies)

Urban (33)

Undeveloped (9)

Land-use studies of shallow ground water

Mixed land use (92)

Major aquifer studies

Bed-Sediment Sampling Sites

Agricultural (242 sites)

Urban (127)

Undeveloped (316)

Mixed (367)

Watershed land use

Stream-Water Sampling Sites

Agricultural (83 sites)

Urban (30)

Undeveloped (8)

Mixed (65)

Watershed land use

NAWQA Study Unit

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3_03_sites

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NAWQA’s Approach to Pesticide Assessment 37
Chemical Analyses

NAWQA chemical analyses encompassed the most complete range of pesticides 

ever measured in a single assessment, and 

included many of the Nation’s most heavily 

used pesticides (Appendix 1). Most NAWQA 

water samples were analyzed for 75 pesticides 

and 8 pesticide degradates, which during the 

study period accounted for about 78 percent of 

the Nation’s agricultural use of conventional 

pesticides, by weight of active ingredient 

(fig. 3–4). The analytical strategy resulted in 

relatively thorough coverage of the major herbicides and insecticides in use for agricultural 

purposes during the period of study, including 20 

of the 25 most heavily used herbicides and 16 of 

the 25 most heavily used insecticides (fig. 3–5). 

All water samples were filtered prior to analysis. 

As a result, reported pesticide concentrations, 

particularly in stream water, may underestimate 

concentrations of some compounds that have 

Figure 3–4. The NAWQA analytical strategy 

for water samples included 75 pesticides that 

accounted for about 78 percent of total national 

agricultural use during the study period, with 

relatively thorough coverage of the major herbicides 

and insecticides, but sparse coverage of fungicides, 

fumigants, and nematicides. (Use estimates are for 

1997, as reported by Gianessi and Marcelli, 2000.)

0 100 200 300 400 500

Annual agricultural use, in millions of pounds

Herbicides

Insecticides

Fungicides

Analyzed by NAWQA

Not analyzed

NAWQA pesticide analyses

in relation to use

Figure 3–5. The NAWQA analytical strategy for water samples included 20 of the 25 most heavily used herbicides 

and 16 of the 25 most heavily used insecticides during the study period. More than half of these 36 high-use pesticides 

included in NAWQA analyses (dark bars) were among those most frequently detected in water (bold type), as described 

in Chapter 4. Some high-use pesticides, such as glyphosate, were omitted because no suitable analytical method was 

available, or because of budget constraints. (Use estimates are for 1997, as reported by Gianessi and Marcelli, 2000.)

Herbicides

High-use pesticides analyzed by NAWQA

Insecticides

Chlorpyrifos

Terbufos

Parathion-methyl

Malathion

Carbaryl

Aldicarb

Carbofuran

Phorate

Cryolite

Propargite

Acephate

Azinphos-methyl

Methomyl

Dimethoate

Endosulfan

Phosmet

Disulfoton

Permethrin

Ethoprop

Methamidophos

Oxamyl

Diazinon

Profenofos

Thiodicarb

Dicofol

0 20 40 60 80 0 4 8 12 16

Annual agricultural use, in millions of pounds Annual agricultural use, in millions of pounds

Atrazine

Metolachlor

2,4-D

Glyphosate

Acetochlor

Pendimethalin

Trifluralin

Cyanazine

Alachlor

Dicamba

EPTC

Propanil

Bentazon

Paraquat

Dimethenamid

MCPA

Fluometuron

Simazine

MSMA

Diuron

Molinate

Metribuzin

Bromoxynil

Clomazone

Norflurazon

Darkened bars indicate

analysis by NAWQA

Pesticides in bold type 

were among the most 

frequently detected in 

water

38 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 3

strong affinities for suspended particles because 

of removal by filtration.

Historically used organochlorine insecticides (such as DDT) and selected degradates 

and by-products were analyzed in bed sediment 

and fish tissue—environmental media in which 

they continue to persist long after the uses of 

the parent compounds were discontinued in the 

United States. Bed-sediment and fish-tissue 

samples were analyzed for up to 32 pesticide 

compounds, consisting of 19 parent pesticides 

and 13 degradates and by-products. Results for 

organochlorines are sometimes described for 

pesticide groups (such as “total DDT”) because 

all compounds in the group are derived either 

from common parent pesticides (for example, 

the isomers p,p'-DDD and p,p'-DDE are degradates of p,p'-DDT), or from indistinguishable 

pesticide products (for example, dieldrin may 

originate from application of dieldrin or or as a 

degradate of aldrin). Pesticide groups are identified and defined in Appendix 1B. Together, the 

organochlorine pesticides examined by NAWQA 

account for more than 90 percent of the Nation’s 

historical use (by weight) of organochlorine pesticides in agriculture.

Some pesticides and degradates were 

assessed only in selected study areas or for limited periods of time. For example, several degradates of alachlor and metolachlor were measured 

at selected sites in different parts of the Nation, 

including streams draining parts of the Great 

and Little Miami River Basins, where use of the 

parent pesticides was high. Some pesticides were 

included in newly developed analytical methods 

late in the study period and were examined only 

in selected high-use areas, as was the case for 

fipronil in the Acadian–Pontchartrain Drainages. 

These and other pesticides with limited data are 

described for selected case studies in this report, 

but are not included in the national analysis. 

Pesticides Not Assessed

Many potentially important pesticides and 

degradates were not assessed because of limited 

analytical methods or budget constraints. For 

example, glyphosate, the pesticide that ranked 

fourth among the top 10 herbicides used in 1997, 

was not routinely analyzed because its chemical 

structure and properties require analytical methods that are different from NAWQA methods, 

which were designed to cost-effectively measure 

large suites of compounds simultaneously. As 

a result, a separate method was developed for 

glyphosate and it was added as an analyte in 

selected studies late in the study period. Similarly, cryolite—an inorganic insecticide used 

on grapes and that ranked ninth among the top 

10 insecticides used in 1997—was also not 

routinely analyzed in NAWQA samples. Of all 

the fungicides used, only chlorothalonil (ranked 

highest in use among fungicides) was included, 

and there were no nationally consistent analyses 

of any fumigants or nematicides. Other pesticide 

compounds that were not assessed by NAWQA 

include other inorganic pesticides (such as 

sulfur and copper), oil, biological pesticides, and 

numerous pesticide degradates, manufacturing 

by-products, and adjuvants. 

Although NAWQA included the broadest 

and most complete range of pesticides measured 

in a single national assessment, it still must be 

considered selective. Consequently, NAWQA 

results should be expected to underestimate the 

overall occurrence of pesticides and degradates 

in many of the hydrologic systems that were 

studied.

NAWQA’s Approach to Pesticide Assessment 39

Chapter 4

Pesticides were found at detectable concentrations in streams 

and ground water within most areas sampled that have substan-

tial agricultural or urban land uses. Pesticides were detected in 

almost every water sample from streams, but were less com-

mon in ground water. Organochlorine pesticide compounds 

were detectable in fish andwom most streams, despite the fact 

that most of the parent pesticides have not been used in the 

United States for years. Although more than 100 pesticide 

compounds were analyzed in water, fish, or sediment, less than 

40 of these compounds accounted for most of the detections. 

The distributions of the most prevalent pesticides in streams 

and ground water largely follow geographic patterns in land 

use and associated present or past pesticide use. 

This chapter summarizes NAWQA 

results from 1992 to 2001, 

focusing on the pesticides that 

were most frequently detected 

in streams and ground water. 

Pesticide detections are assessed 

in relation to land use, pesticide 

use, hydrologic settings, and 

the properties of the pesticides 

themselves. More detailed 

examination of seasonal patterns, 

pesticide mixtures, and degradates 

is included in Chapter 5, and a 

screening-level assessment of the 

potential significance of pesticides 

to human health, aquatic life, and 

wildlife is provided in Chapter 6.

4

Occurrence and Distribution in Streams 

and Ground Water

Photograph by Eric Hatch Photography, Loveland, Ohio
Overview of Pesticide Occurrence

NAWQA results show that pesticides 

occurred at detectable concentrations in streams 

and ground water within most areas that have 

substantial agricultural or urban land uses. These 

findings build upon an extensive body of previous research, demonstrating that pesticides and 

their degradates are present in ground water 

(for example, Barbash and Resek, 1996; Kolpin 

and others, 1996), surface waters (for example, 

Larson and others, 1997; Battaglin and others, 

2003), and stream sediments and aquatic biota 

(for example, Nowell and others, 1999; Seiler 

and others, 2003) in a wide variety of hydrologic, ecological, and land-use settings across the 

United States.

In streams sampled by NAWQA, at least one 

pesticide or degradate was detected more than 90 

percent of the time in water, in more than 80 percent of fish samples, and in more than 50 percent 

of bed-sediment samples collected during 1992–

2001 (fig. 4–1). Pesticides analyzed in water 

were primarily those that are currently used, 

whereas those analyzed in fish and sediment were 

predominantly pesticides (or their degradates and 

by-products) that are no longer used in the United 

States, such as DDT and other organochlorine 

pesticides. Detections in stream water were 

evaluated on a time-weighted basis and results 

are expressed as the percentage of time that concentrations were detectable. For both fish tissue 

and bed sediment, one sample was analyzed per 

site and detections are expressed as a percentage 

of samples or sites analyzed. Detectable concentrations occurred in water more than 90 percent 

of the time for streams draining watersheds with 

agricultural, urban, and mixed land use. Similarly, organochlorine pesticide compounds were 

detected in more than 90 percent of fish-tissue 

samples and in more than 50 percent of bed-sediment samples from streams in watersheds with 

agricultural, urban, and mixed land use. In water, 

fish tissue, and bed sediment, detections were the 

least frequent, but not absent, for streams draining undeveloped watersheds—where pesticide 

use is lowest.

Pesticides were detected distinctly less often 

in ground water than in streams (fig. 4–1). Detections in ground water are based on one sample 

per well. Streams are generally more vulnerable 

to contamination than ground water because of 

the direct and relatively rapid overland transport 

of pesticides that occurs with surface runoff (see 

Chapter 2). Ground water in most areas is less 

vulnerable because water infiltrates the land 

surface and moves slowly through soil and aquifer materials before reaching most wells. This 

extended travel time allows more opportunities 

for the concentrations of pesticides in water to 

Figure 4–1. Considering all streams sampled across all land uses, one or more pesticides or their degradates were detectable 

more than 90 percent of the time in water, and were detected in more than 80 percent of fish samples and in more than 50 

percent of bed-sediment samples. Less than half of all ground-water samples contained one or more detected pesticides or their 

degradates, with the most frequent detections occurring in shallow ground water beneath agricultural and urban areas.

Agricultural areas

Pesticides in water

(most were used during the study period)

Organochlorine compounds in fish and sediment

(most are no longer used in the U.S.)

Urban areas

Undeveloped areas

Mixed land uses

Stream water

Shallow ground water

Stream water

Shallow ground water

Stream water

Shallow ground water

Stream water

Major aquifers

0 25 50 75 100

Percentage of time (streams) or samples

(ground water) with one or more 

detectable pesticides or degradates

Overview of pesticide occurrence

Agricultural areas

Urban areas

Undeveloped areas

Mixed land uses

Fish tissue

Bed sediment

Fish tissue

Bed sediment

Fish tissue

Bed sediment

Fish tissue

Bed sediment

0 25 50 75 100

Percentage of samples with one or more

detectable pesticides or degradates

4_01freqexOVsummary

42 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 4

be reduced from the combined action of sorption, dispersion, dilution, and transformation. The 

slow movement of water and solutes through the 

subsurface, however, also makes contamination 

of ground water more difficult to reverse once it 

occurs. The highest frequencies of detection were 

in shallow ground water beneath agricultural and 

urban areas, where almost 60 percent of samples 

had detections of one or more pesticides or degradates. The lowest frequencies of detection were 

found in shallow ground water beneath undeveloped areas and in deeper ground water in major 

aquifers. Samples from major aquifers generally 

represent older ground water that originated as 

recharge in areas of mixed land use, sometimes 

before the current land uses were present.

Pesticides Detected Most Frequently

About 40 pesticide compounds, of the more 

than 100 examined by NAWQA, accounted for 

most detections in water, fish, or bed sediment. 

Understanding the occurrence and distribution of 

these most prevalent pesticides—both spatially 

and temporally—in relation to their use and 

properties, land use, and hydrologic settings is 

critical for evaluating the potential significance of 

pesticides to water quality. 

Water

Twenty-five pesticide compounds, including 

24 pesticides and 1 degradate, were each detected 

more than 10 percent of the time in streams in 

agricultural, urban, or mixed-land-use settings, or 

in more than 2 percent of wells in agricultural or 

urban settings (fig. 4–2). These 25 pesticide compounds include 11 of the herbicides used most 

heavily in agriculture during the study period 

(plus the atrazine degradate, deethylatrazine)—

hereinafter referred to collectively as agricultural 

herbicides; 7 herbicides used extensively (though 

not exclusively) for nonagricultural purposes—

referred to as urban herbicides; and 6 insecticides 

used in both agricultural and urban settings, but 

most intensively in urban settings (fig. 4–3).

The broad patterns of pesticide occurrence 

in streams generally corresponded to land use 

and pesticide use. For example, major agricultural herbicides, such as atrazine and metolachlor, 

were found most often in agricultural settings, 

whereas herbicides frequently used in urban 

areas, such as simazine and prometon, were 

found most often in urban settings. Urban herbicides also were detected in some agricultural 

areas, either because of agricultural uses (such 

as for simazine), or their use for nonagricultural 

weed control (such as for prometon). Insecticides 

were generally found most often in urban settings 

where, with the exception of carbofuran, they are 

used more intensively than in most agricultural 

settings. Patterns of detection in ground water 

also generally corresponded with patterns of land 

use, although not as closely as for streams. For 

example, atrazine and metolachlor were among 

the pesticides detected most frequently in both 

streams and shallow ground water in agricultural 

areas.

Pesticides were detected least often in 

streams and shallow ground water in undeveloped areas. The occurrence of atrazine, deethylatrazine, and metolachlor in undeveloped streams 

was likely caused by one or more of several 

factors: (1) most undeveloped watersheds include 

small areas of agricultural or urban land; (2) pesticides are used in many undeveloped areas for a 

variety of purposes, such as pest control in forest 

lands or weed control along utility and roadside 

rights-of-way; and (3) pesticides can be transported in the atmosphere from other areas. 

Not surprisingly, the pesticides that were 

most commonly detected in streams draining watersheds with mixed land use reflected 

multiple sources from agricultural and urban 

applications. The overall frequency of pesticide occurrence in mixed-land-use streams was 

similar to those observed in both agricultural and 

urban streams. Likewise, the pesticides detected 

in major aquifers indicate the influences of both 

agricultural and urban sources, but overall detection frequencies were lower in major aquifers 

than in shallow ground water in agricultural and 

urban areas.

The pesticide detected most frequently in streams and 

ground water was atrazine, an herbicide used to control 

weeds in corn fields.

Occurrence and Distribution in Streams and Ground Water 43
Figure 4–2. Consistent with their patterns of use during the study period, agricultural herbicides—most commonly atrazine (and its 

degradate deethylatrazine), metolachlor, cyanazine, alachlor, and acetochlor—were detected more frequently in agricultural areas than 

in urban areas; urban herbicides were found most often in urban areas; and most insecticides, such as diazinon and carbaryl, were 

detected more frequently in urban streams than in agricultural streams. Two different detection levels are used in this analysis. The dark 

portion of each bar indicates detections at concentrations greater than or equal to 0.1 μg/L , the light portion indicates detections less 

than 0.1μg/L*, and the end of each bar is the total for all detections.

*The pesticides 2,4-D, bentazon, bromacil, carbaryl, diuron, and norflurazon could not be detected reliably at concentrations less than 0.1 μg/L; 

consequently, the reported frequencies below this level for these compounds are minimum estimates.

Agricultural

Atrazine

Deethylatrazine

Metolachlor

Cyanazine

Alachlor

Acetochlor

Metribuzin

*Bentazon

EPTC

Trifluralin

Molinate

*Norflurazon

Simazine

Prometon

Tebuthiuron

*2,4-D

*Diuron

Dacthal

*Bromacil

Diazinon

Chlorpyrifos

Carbofuran

*Carbaryl

Malathion

Dieldrin

Atrazine

Deethylatrazine

Metolachlor

Cyanazine

Alachlor

Acetochlor

Metribuzin

*Bentazon

EPTC

Trifluralin

Molinate

*Norflurazon

Simazine

Prometon

Tebuthiuron

*2,4-D

*Diuron

Dacthal

*Bromacil

Diazinon

Chlorpyrifos

Carbofuran

*Carbaryl

Malathion

Dieldrin

Urban

Percentage of samples with detections

Streams

Pesticides detected most frequently in water

Percentage of time detected

Ground Water

Undeveloped Mixed land use

Agricultural Urban Undeveloped Major aquifers

0 25 50 75 100 0 25 50 75 100 0 25 50 75 100 0 25 50 75 100

0 25 50 75 100 0 25 50 75 100 0 25 50 75 100 0 25 50 75 100

Insecticides

Urban

herbicides

Agricultural

herbicides Insecticides

Urban

herbicides

Agricultural

herbicides

4_02ocur

44 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 4

Understanding the Occurrence Assessment

The overview described in this chapter serves as a broad first step 

toward understanding the distribution and importance of pesticides that 

were detected in streams and ground water. As explained in Chapter 3, 

the NAWQA assessment did not include samples from all parts of the 

Nation or include all pesticides currently used. To provide a national perspective, the occurrence of pesticide compounds in streams and ground 

water is summarized by land use, environmental medium and component 

of the hydrologic system sampled, and detection level—all of which 

have important influences on how results are interpreted.

Land use—Grouping results by land use follows the NAWQA design 

by combining data from sites expected to have similar influences 

of land use on water quality (see Chapter 3). Within each general 

land-use setting, however, there can be substantial variability 

among sampling sites in specific land-use conditions and pesticide 

use, as well as hydrologic settings. A pesticide that is common in 

agricultural streams as a national group, such as a corn herbicide, 

may never occur in some particular agricultural streams, whereas 

another pesticide that is uncommon nationally, such as a rice herbicide, may be frequently detected in a few particular streams.

Media and hydrologic component sampled—Grouping results by 

the different environmental media that were sampled from streams 

clearly separates organochlorine compounds—which were derived 

primarily from past use and which, because of their hydrophobic 

nature, were assessed by their occurrence in fish tissue and bed 

sediment—from predominantly water-soluble pesticides, most of 

which were in use during the study period and were measured 

in water. Grouping results by hydrologic component further distinguishes between streams and ground water for analysis of 

pesticides in water. The occurrence and concentration results for 

stream water, unless noted otherwise, were evaluated on a timeweighted basis for each site to eliminate biases caused by more 

frequent sampling during high-use seasons. Ground-water results 

are based on one sample per well, and bed-sediment and fishtissue results are based on one composite sample per site.

Detection level—Analyses of pesticide occurrence in this chapter 

are based on two different detection levels: (1) detection at any 

concentration—as low as 0.001 μg/L in water—referred to as total 

detection frequency, and (2) detections greater than or equal to a 

common detection level for all compounds in a particular medium—

0.1 μg/L for water, 5 micrograms per kilogram (μg/kg) wet weight for 

fish tissue, and 2 μg/kg dry weight for bed sediment. Two detection 

levels are necessary for certain data analyses because variations 

in analytical sensitivity result in differences in minimum detectable 

concentrations among different compounds. Consequently, direct 

comparisons of detection frequencies among compounds should 

be based on a common detection level. For example, of the 25 pesticide compounds most frequently detected in water, 2,4-D, bentazon, 

bromacil, carbaryl, diuron, and norflurazon could not be reliably 

detected at concentrations less than about 0.1 μg/L, whereas the 

other 19 compounds were detectable at levels substantially less 

than 0.1 μg/L. The reported total detection frequencies of the 6 

less-detectable pesticides are thus underestimates of occurrence 

compared with the total frequencies for the other 19 compounds. 

Variations in detection sensitivity must be carefully considered 

when interpreting data on occurrence—the absence of detections 

does not indicate with certainty that pesticides are not present.

•

•

•

Figure 4–3. The pesticides detected most frequently 

in water include 11 of the herbicides used most heavily 

for agriculture during the study period, 7 herbicides 

used extensively for nonagricultural purposes (mostly 

in urban areas, but with some agricultural applications), 

and 6 insecticides. (Agricultural-use estimates are from 

Gianessi and Marcelli [2000] for 1997; nonagriculturaluse estimates are from Kiely and others [2004] for 1999, 

but were available only for 2,4-D, carbaryl, chlorpyrifos, 

dacthal, diazinon, diuron, and malathion—indicated by 

bold type.)

Malathion

Dieldrin

Carbaryl

Carbofuran

Chlorpyrifos

Diazinon

Bromacil

Diuron

2,4-D

Tebuthiuron

Prometon

Simazine

Molinate

Norflurazon

Dacthal

Trifluralin

EPTC

Bentazon

Metribuzin

Acetochlor

Alachlor

Cyanazine

Metolachlor

Atrazine

Insecticides Urban

herbicides

Agricultural

herbicides

no longer used

0 20 40

Annual use, in millions of pounds

60 80

Agricultural use

Nonagricultural use

4_03use

Use of frequently detected 

pesticides

Occurrence and Distribution in Streams and Ground Water 45
Fish Tissue and Bed Sediment

Thirteen organochlorine pesticide compounds, including historically used parent 

pesticides and their degradates and by-products, 

were each found in more than 10 percent of fish 

or bed-sediment samples from streams draining 

watersheds with either agricultural, urban, or 

mixed land use. Figure 4–4 summarizes findings 

for these 13 compounds, as well as for 2 additional compounds derived from DDT use—

o,p'-DDT and its degradate, o,p'-DDE, which 

were detected less frequently than the others. The 

fish and bed-sediment data for organochlorine 

compounds provide complementary types of 

information for understanding the presence of 

these compounds in streams.

The 15 organochlorine pesticide compounds 

included in figure 4–4 are derived from 8 parent 

pesticides. The parent pesticides applied were the 

insecticides DDT, DDD (also known as TDE), 

dieldrin, aldrin, chlordane, and heptachlor—each 

of which had all agricultural and nonagricultural uses cancelled by 1988 or earlier, and the 

fungicides hexachlorobenzene and pentachlorophenol—most uses of which were discontinued 

by the mid 1980s or before. DDT and chlordane 

were applied as technical mixtures containing 

the parent pesticides and other compounds. For 

example, technical DDT was typically composed 

cis-Chlordane

trans-Chlordane

cis-Nonachlor

trans-Nonachlor

Oxychlordane

Dieldrin

Heptachlor epoxide

Pentachloroanisole

Hexachlorobenzene

p,p'-DDE

p,p'-DDD

p,p'-DDT

o,p'-DDE

o,p'-DDD

o,p'-DDT

0 25 50 75 100 0 25 50 75 100 0 25 50 75 100 0 25 50 75 100

0 25 50 75 100 0 25 50 75 100 0 25 50 75 100 0 25 50 75 100

Agricultural Urban

Urban

Fish Tissue

Undeveloped Mixed land use

Agricultural Undeveloped Mixed land use

Bed Sediment

Percentage of samples with detections

Percentage of samples with detections

Total DDT

cis-Chlordane

trans-Chlordane

cis-Nonachlor

trans-Nonachlor

Oxychlordane

Dieldrin

Heptachlor epoxide

Pentachloroanisole

Hexachlorobenzene

p,p'-DDE

p,p'-DDD

p,p'-DDT

o,p'-DDE

o,p'-DDD

o,p'-DDT

Total DDT

Total

Chlordane

Total

Chlordane

Organochlorine compounds detected most frequently

Figure 4–4. Historically used organochlorine pesticides and their degradates were generally detected 

more frequently in whole fish and bed sediment in urban streams than in agricultural streams, thus matching 

the pattern found for currently used insecticides in water. DDT and chlordane compounds, as well as 

dieldrin, were relatively widespread. The dark portion of each bar indicates detections at concentrations 

greater than or equal to 5 μg/kg (wet weight) for fish tissue and 2 μg/kg (dry weight) for bed sediment, the 

light portion indicates detections at concentrations less than these levels, and the end of each bar is the 

total for all detections.

46 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 4

of about 80 percent p,p'-DDT (the active ingredient) and 20 percent o,p'-DDT (o,p' and p,p'

indicate different isomers of DDT). Furthermore, 

in addition to being applied as pesticides, DDD 

and dieldrin are also formed in the environment 

from the transformation of DDT and aldrin, 

respectively. Dieldrin that originated from the 

application of aldrin could not be distinguished 

from dieldrin applied as dieldrin. Thus, for the 

purposes of certain data analyses, parent pesticide compounds were grouped together with 

their corresponding degradates and by-products, 

reflecting their common or indistinguishable 

origins. 

Six compounds were analyzed in the DDT 

group (the p,p' and o,p' isomers of DDT, DDE, 

and DDD). The sum of the concentrations of 

these compounds is referred to as the total DDT 

concentration. Five compounds were analyzed 

in the chlordane group, with total chlordane 

concentration calculated as the sum of concentrations of the cis and trans isomers of chlordane 

and nonachlor, plus the chlordane degradate 

oxychlordane. Additional individual compounds 

frequently found in streams included dieldrin; 

pentachloroanisole, a degradate of pentachlorophenol; heptachlor epoxide, a degradate of heptachlor; and hexachlorobenzene. Historically, DDT, 

DDD, aldrin, and dieldrin were used widely 

in both agricultural and urban areas, whereas 

chlordane use for urban applications was greater 

than its agricultural use. As shown in figure 4–5, 

the agricultural uses of DDT plus DDD, and of 

dieldrin plus aldrin, were higher than the uses of 

heptachlor and chlordane. Most organochlorine 

insecticides had their agricultural uses discontinued in the 1970s, whereas some urban applications (including termite control) were permitted 

until the late 1980s. 

Results for fish tissue and bed sediment 

show generally similar patterns of detection 

among the organochlorine pesticide compounds 

(fig. 4–4), but detections were more frequent in 

fish tissue because these compounds typically 

accumulate to higher concentrations in biological tissues (wet-weight concentrations) than in 

sediment (dry-weight concentrations). Patterns 

of occurrence of organochlorine compounds in 

fish and bed sediment generally match the patterns in relation to land use that are evident for 

currently used insecticides in water. Frequencies 

of detection were higher for most organochlorine pesticide compounds in urban streams than 

in agricultural streams. The most frequently 

detected compounds were those composing the 

DDT group, the chlordane group, and dieldrin.

Streams with undeveloped watersheds 

had the lowest frequencies of detection of 

organochlorine compounds in either fish or bed 

sediment, yet more than half of the fish-tissue 

samples from these streams had detectable levels 

of p,p′-DDE, a principal degradate of DDT. 

The frequent presence of p,p′-DDE in fish from 

undeveloped streams may be explained by factors 

similar to those believed to result in the presence of currently used pesticides in water from 

undeveloped streams: (1) past use in small areas 

of developed land within their watersheds, (2) 

past use for control of insects in undeveloped 

areas (such as for forest management), and (3) 

atmospheric transport from other areas.

Fish and bed sediment from streams draining watersheds with mixed land use had frequencies of detection of DDT, dieldrin, and chlordane 

that generally reflected a mixture of agricultural 

and urban influences, and were higher than in 

undeveloped streams (fig. 4–4).

Figure 4–5. In 1966, the agricultural uses of DDT 

plus DDD (mostly as DDT), and of dieldrin plus aldrin 

(mostly as aldrin), were greater than the presentday agricultural use of any individual insecticide. 

Heptachlor and chlordane had much lower agricultural use. (Use estimates are from Eichers and others, 

1970.)

Chlordane

Heptachlor

Dieldrin and aldrin

DDT and DDD

0 5 10 15 20

Annual agricultural use, in millions of pounds

25 30 35

Historical use of organochlorine

pesticides

Occurrence and Distribution in Streams and Ground Water 47
Influence of Land Use

The preceding overviews of detection 

frequencies clearly show that pesticide occurrence in streams and ground water is strongly 

influenced by land use and associated pesticide 

use. These relations are explored in more detail 

below, focusing on the occurrence of some of the 

pesticides that were detected most frequently in 

agricultural and urban settings. 

Agricultural Areas

The pesticides detected most often in water 

from agricultural streams (fig. 4–2) were among 

the agricultural herbicides used most heavily during the study period (figs. 4–3 and 3–5). The top 

five, from highest to lowest frequency of detection at concentrations at or above 0.1 μg/L, were 

atrazine (ranked 1st in national agricultural use; 

deethylatrazine was also frequently detected), 

metolachlor (2nd in use), cyanazine (8th in use), 

2,4-D (3rd in use), and simazine (18th in use). 

Prometon was detected frequently at low levels 

(rarely at concentrations greater than 0.1 μg/L) 

and ranked only behind atrazine, metolachlor, 

simazine, and cyanazine in total detection 

frequency. Prometon is not registered for use on 

crops, but is used for weed control around fences, 

buildings, and roads within agricultural areas. 

As with streams, the pesticides most commonly found in shallow ground water within 

agricultural areas were atrazine (and deethylatrazine) and metolachlor, the two herbicides used 

most heavily for agriculture during the study 

period. Although atrazine and metolachlor had 

about the same total use, atrazine and deethylatrazine were found in ground water more than 

twice as often as metolachlor, probably because 

atrazine is considerably more persistent than 

metolachlor (as discussed in more detail later 

in this chapter). Deethylatrazine was detected 

in ground water about as frequently as atrazine, 

whereas in streams it was found less often than 

atrazine and usually at lower levels. The greater 

proportional occurrence of deethylatrazine in 

ground water reflects the greater opportunity 

for atrazine degradation over the longer periods 

of time that water in the subsurface spends in 

contact with microbes, especially in the soil zone 

(as discussed further in Chapter 5). Cyanazine, 

alachlor, and acetochlor—which are used on corn 

and other crops, but in less than half the amounts 

of atrazine and metolachlor—were seldom 

detected in ground water, most likely because 

of their lower use and relatively low persistence 

(Appendix 2). In contrast, simazine and prometon were among the pesticides found most often 

in ground water, despite even lower agricultural 

use than cyanazine, alachlor, or acetochlor. Simazine and prometon are more persistent in soil 

than these other herbicides, and thus have greater 

opportunities for transport to ground water.

Currently used insecticides were found less 

frequently than herbicides in most agricultural 

streams and were rarely found in ground water. 

This finding results from their relatively low 

application rates in most agricultural settings, 

compared with herbicides (fig. 3–5), and their 

generally lower mobility and persistence in the 

environment (Appendix 2). The insecticide used 

most heavily for agricultural purposes during the 

study period was chlorpyrifos. Yet, annual use 

of chlorpyrifos was only about 20 percent that of 

atrazine use, and chlorpyrifos is also less mobile 

in the hydrologic system. Although the annual 

agricultural use of each of the other four major 

insecticides examined—diazinon, carbofuran, 

carbaryl, and malathion—was less than half that 

of chlorpyrifos during the study period, the total 

detection frequencies of all five insecticides in 

agricultural streams were notably similar.

Historically used organochlorine pesticides 

and their degradates and by-products remained a 

common occurrence in fish and bed sediment in 

agricultural streams, although most were detected 

less frequently in samples from agricultural 

streams than from urban streams—especially in 

sediment (fig. 4–4). The compounds found most 

commonly in agricultural streams were those 

in the DDT group, followed by dieldrin and the 

chlordane group. Relative frequencies of detection corresponded to their rankings of historical 

use in agriculture (fig. 4–5). The frequency of 

occurrence of compounds in the chlordane group 

in agricultural streams was higher than expected 

from its low historical agricultural use compared 

with DDT plus DDD and dieldrin plus aldrin—

possibly because of extensive nonagricultural 

applications in agricultural areas (such as termite 

control). In addition, chlordane was a minor 

component (10–20 percent) of technical-grade 

heptachlor, which was also used extensively both 

in agriculture and as a termiticide (IARC, 2001). 

Although these broad patterns in pesticide 

occurrence across all agricultural areas that were 

sampled provide a useful national perspective, 

the aggregated results obscure many substantial 

differences among different agricultural settings 

in the types and levels of pesticides that were 

detected. The many diverse agricultural settings 

48 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 4

of the United States that were sampled—each 

with its own unique combination of climate, 

crops, and pests—have distinctive patterns in 

pesticide use that resulted in different patterns of 

pesticide occurrence. These patterns of occurrence are complex because of the wide ranges of 

different use practices, pesticide properties, and 

hydrologic processes that govern the sources, 

movement, and persistence of pesticides in 

streams and ground water. 

Comparisons of patterns in pesticide occurrence among three of the Nation’s major cropgroup settings illustrate the variability among 

settings. Classification of the Nation’s agricultural areas for the NAWQA water-quality studies 

identified 21 major crop-group settings of varying areal extent (Gilliom and Thelin, 1997). This 

classification is based on combinations of one to 

three crops that account for most of the harvested 

acreage in each of the Nation’s counties. Three 

crop-group settings were selected as examples for 

comparison in this report: “corn and soybeans,” 

“wheat and alfalfa,” and “rice.” Each crop-group 

setting has a different geographic distribution 

and extent (fig. 4–6). Other crops are also present 

to varying degrees in each of the three settings; 

thus, the estimated use of 

a pesticide in a particular crop setting may also 

include its estimated use 

for other crops in the same 

area. 

Estimates of pesticide 

use intensity, expressed as 

an annual average rate of 

application on all cropland 

in each crop-group setting, 

show clear differences 

between the settings 

(fig. 4–7). Overall rates 

of use were highest in the 

corn-and-soybeans setting 

and lowest in the wheatand-alfalfa setting. Use in 

each setting is dominated 

by the particular herbicides 

and insecticides needed to 

control the pests specific 

to the crops grown in that 

setting. For example, 

atrazine and metolachlor 

dominated herbicide use 

in the corn-and-soybeans 

setting; 2,4-D was the 

top herbicide used in the 

wheat-and-alfalfa setting 

Figure 4–7. Different pesticides dominated use in each of the crop-group settings, as illustrated by the 

estimated 1997 agricultural use. The herbicides atrazine, metolachlor, acetochlor, and cyanazine were 

the most intensively used pesticides in the corn-and-soybeans setting; molinate, 2,4-D, and several 

insecticides were most intensively used in the rice setting; and 2,4-D and chlorpyrifos were the most 

intensively used pesticides in the wheat-and-alfalfa setting, where overall use was least.

Dieldrin

Malathion

Carbaryl

Carbofuran

Chloryprifos

Diazinon

Bromacil

Dacthal

Diuron

2,4-D

Tebuthiuron

Prometon

Simazine

Norflurazon

Molinate

Trifluralin

EPTC

Bentazon

Metribuzin

Acetochlor

Alachlor

Cyanazine

Metolachlor

Atrazine

Insecticides

Urban

herbicides

Agricultural

herbicides

0 100 200 300 400 0 100 200 300 400 0 100 200 300 400

Annual agricultural use intensity, in pounds per square mile

Corn and soybeans Wheat and alfalfa Rice

no use data no use data no use data

no longer used no longer used no longer used

Pesticide use by crop setting

Wheat and alfalfa

Rice

Corn and soybeans 

Stream

site

Ground-water

study

Crop-group setting

Distribution of study sites by crop setting

Figure 4–6. The distributions of crop-group settings 

and study sites for the corn-and-soybeans, wheat-andalfalfa, and rice crop groups show distinct differences 

in the locations and extent of the three agricultural 

settings. (Crop groups are from Gilliom and Thelin, 

1997.)

Occurrence and Distribution in Streams and Ground Water 49
(but at less than half the intensity of atrazine 

and metolachlor use in the corn-and-soybeans 

setting); and molinate was the top herbicide used 

in the rice setting (with use intensity that was 

70 percent greater than that of either atrazine or 

metolachlor in the corn-and-soybeans setting).

The occurrence of pesticides in streams and 

ground water within these three crop settings 

(fig. 4–8) corresponded to the estimated agricultural-use patterns in many respects (fig. 4–7), but 

also showed some unexpected results, as summarized below. These examples of results for specific crop-group settings illustrate both the degree 

of predictability and the complexity of pesticide 

occurrence and transport in the hydrologic 

system. Each crop setting has unique characteristics, and each specific study area within a crop 

setting is unique as well, resulting in variability 

within crop-group settings as well as among 

them. Nonetheless, organizing the assessment 

of pesticides by crop-group setting can help to 

link the occurrence of pesticides in streams and 

ground water with specific management practices 

and can provide the foundation for customizing 

pesticide management to individual settings.

Expected results:

Corn and soybeans—The two herbicides used most heavily for corn 

and soybeans—atrazine and metolachlor—were those detected most 

frequently in streams and ground water. In addition, deethylatrazine was 

detected at about the same frequency as atrazine in both streams and 

ground water within this setting. Chlorpyrifos was both the most frequently detected and the most heavily used insecticide.

Wheat and alfalfa—Overall detection frequencies were low in the wheatand-alfalfa setting, consistent with relatively low pesticide use. The herbicide used most heavily in the wheat-and-alfalfa setting, 2,4-D, was one 

of the most frequently detected pesticides in streams at concentrations at 

or above 0.1 μg/L.

Rice—Molinate, the herbicide used most heavily on rice, was among 

those detected most frequently in streams. Detections of molinate were 

far more frequent in the rice setting than in the other agricultural settings. 

The insecticide used most intensively on rice, carbofuran, was the one 

detected most frequently at or above 0.1 μg/L in both streams and ground 

water. Carbofuran and the other four insecticides used mostly in the rice 

setting were detected more frequently in the rice setting than in the other 

two crop-group settings, where their use was less intensive.

Unexpected results:

Corn and soybeans—Simazine and prometon were found more frequently than was expected from their low agricultural use, indicating 

relatively substantial use of these herbicides for noncrop purposes within 

agricultural areas (although most concentrations were low). 

Wheat and alfalfa—Atrazine and prometon were the herbicides detected 

most frequently in streams and ground water, despite little (atrazine) or 

no (prometon) agricultural use on either wheat or alfalfa (although most 

concentrations were low).

Rice—Low-use pesticides, including atrazine (and deethylatrazine), 

metolachlor, and tebuthiuron, were frequently detected, probably because 

of noncrop uses within this setting. Bentazon was frequently detected in 

streams and particularly in ground water. Bentazon was detected most 

frequently in the rice-growing area of California, where it was used heavily until it was banned in 1989.

•

•

•

•

•

•

Photographs by Don Brennemen, 

University of Minnesota 

Agricultural Extension Service 

(middle), and © 2003 Corbis 

(top).

50 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 4

Figure 4–8. The occurrence of pesticides in streams and ground water sampled within the corn-and-soybeans, 

wheat-and-alfalfa, and rice crop-group settings corresponded to the patterns of estimated agricultural use in 

many respects (see fig. 4–7), but nonagricultural uses also influence occurrence. The dark portion of each bar 

indicates detections at concentrations greater than or equal to 0.1 μg/L , the light portion indicates detections less 

than 0.1μg/L*, and the end of each bar is the total for all detections (see sidebar on p. 45.)

*The pesticides 2,4-D, bentazon, bromacil, carbaryl, diuron, and norflurazon could not be detected reliably at concentrations 

less than 0.1 μg/L, and the reported frequencies below this level for these compounds are minimum estimates.

Atrazine

Deethylatrazine

Metolachlor

Cyanazine

Alachlor

Acetochlor

Metribuzin

*Bentazon

EPTC

Trifluralin

Molinate

*Norflurazon

Simazine

Prometon

Tebuthiuron

*2,4-D

*Diuron

Dacthal

*Bromacil

Diazinon

Chlorpyrifos

Carbofuran

*Carbaryl

Malathion

Dieldrin

Atrazine

Deethylatrazine

Metolachlor

Cyanazine

Alachlor

Acetochlor

Metribuzin

Bentazon

EPTC

Trifluralin

Molinate

*Norflurazon

Simazine

Prometon

Tebuthiuron

*2,4-D

*Diuron

Dacthal

*Bromacil

Diazinon

Chlorpyrifos

Carbofuran

*Carbaryl

Malathion

Dieldrin

0 25 50 75 100 0 25 50 75 100 0 25 50 75 100

0 25 50 75 100 0 25 50 75 100 0 25 50 75 100

Streams

Corn and soybeans

(35 sites)

Wheat and alfalfa

(3 sites)

Rice

(6 sites)

Ground Water

Corn and soybeans

(14 studies)

Wheat and alfalfa

(2 studies)

Rice

(2 studies)

Percentage of samples with detections

Percentage of time detected

Insecticides

Urban

herbicides

Agricultural

herbicides Insecticides

Urban

herbicides

Agricultural

herbicides

*

4_8

Patterns of detections in water by crop setting

Occurrence and Distribution in Streams and Ground Water 51
Urban Areas

The most distinct differences between 

pesticides found in urban and agricultural areas 

were the more frequent detections and higher 

concentrations of insecticides in urban streams, 

and the frequent detections of urban herbicides 

in streams and shallow ground water sampled in 

urban areas.

Diazinon, chlorpyrifos, carbaryl, and 

malathion, which nationally ranked 2nd, 4th, 

8th, and 15th among pesticides in frequencies of 

outdoor applications for home-and-garden use at 

the beginning of the study period (Whitmore and 

others, 1992), accounted for most detections of 

insecticides in urban streams (fig. 4–2). Diazinon 

and carbaryl were by far the most frequently 

detected and were found at frequencies and levels 

comparable to those for the common herbicides. 

The use of diazinon and chlorpyrifos has been 

substantially curtailed since the end of the study 

period, and analysis of recent data for diazinon 

(Chapter 8) shows that concentrations in some 

streams have now declined as well. Historically 

used insecticides also were found most frequently 

in fish and bed sediment from urban streams, 

which had among the highest detection frequencies for chlordane compounds, DDT compounds, 

and dieldrin (fig. 4–4). Urban streams also had 

the highest concentrations of total chlordane and 

dieldrin in both sediment and fish tissue. Chlordane and aldrin were widely used for termite 

control until the mid-to-late 1980s, although 

their agricultural uses were restricted during 

the 1970s. 

Insecticides were seldom detected in ground 

water beneath urban areas (fig. 4–2). The most 

commonly detected insecticide in shallow ground 

water in urban areas, however, was dieldrin, 

which was found in about 5 percent of the wells 

sampled. Although dieldrin is not very mobile 

in water, its environmental persistence and the 

extensive historical use of dieldrin and aldrin 

have apparently combined to yield detectable 

concentrations in some wells 5 to 15 years after 

all uses of dieldrin and aldrin were discontinued. 

The most frequently detected herbicides in 

streams and shallow ground water in urban areas 

were atrazine (and deethylatrazine), simazine, 

prometon, and metolachlor, although metolachlor 

was seldom detected in ground water—probably because of its lower urban use and lower 

persistence compared with the other herbicides. 

Considering only detections at or above 0.1 μg/L, 

however, the herbicides detected most frequently 

in urban streams were diuron (14 percent of the 

time), simazine (14 percent), 2,4-D (11 percent), 

and atrazine (10 percent).

The herbicides found more often in urban 

areas than in most agricultural areas—considering detections at all concentrations—were simazine, prometon, diuron, 2,4-D, tebuthiuron, and 

dacthal. The use of 2,4-D and prometon ranked 

1st and 14th among herbicides in frequency of 

outdoor home-and-garden applications at the 

beginning of the study period (Whitmore and others, 1992). Although 2,4-D, simazine, and diuron 

also ranked 3rd, 18th, and 23rd among herbicides 

in national use for agriculture, no agricultural use 

was reported for prometon or tebuthiuron. 

Pesticides are used extensively in residential 

areas and associated recreational and 

commercial areas, including golf courses.

52 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 4

Methods and Statistics for Assessing Geographic Distributions of 

Pesticides

Consistent measures and scales are used to represent concentration levels appropriate to each 

medium in the comparisons of geographic distributions among pesticides (figs. 4–9 to 4–16): 

For pesticides in stream water, maps in this chapter are based on the time-weighted 95th-percentile 

concentration at each site for the selected year of data, which is the concentration exceeded about 5 

percent of the time, or about 18 days per year (generally not consecutive). Use of the 95th percentile 

for comparisons reduces the influence of different detection levels among compounds because it is 

usually higher than the lowest detectable concentration.

For pesticides in ground water, maps in this chapter are based on the frequency of detections at or 

above 0.01 μg/L within each study area. Evaluation of each of the pesticides using only detections at or 

above the detection level of 0.01 μg/L yields results that are directly comparable among all pesticides 

mapped for ground water. Symbols representing ground-water studies are shown at the centroid of 

each study area.

For total DDT and dieldrin in streams, data for bed sediment are used because fish were not collected 

in all parts of the country. One composite bed-sediment sample was collected at each site—maps are 

based on the concentration in each individual sample.

For all maps, the distribution of agricultural use for each pesticide is shown by a consistent set of 

categories of 1997 use intensity—or historical use intensity for total DDT and dieldrin—so that maps 

can be directly compared among the 10 pesticides. Use was estimated for 1997 by combining the 1997 

state-level use data reported by Gianessi and Marcelli (2000) with county crop acreages from the 1997 

Census of Agriculture (U.S. Department of Agriculture, 1999), using methods described by Thelin and 

Gianessi (2000). Use intensity was mapped for agricultural land using land-cover data from the early 

1990s (Vogelmann and others, 2001) as described by Nakagaki and Wolock (2005). Historical use of 

DDT (including DDD) and dieldrin (including aldrin) was estimated by a similar approach, but using 

regional use estimates for 1966 (Eichers and others, 1970) and 1971 (Andrilenas, 1974), and the 1964 

and 1969 Censuses of Agriculture for crop distributions (Nowell and others, 2006). Use intensity was 

mapped for agricultural land using land-cover data from the mid-1970s (Fegeas and others, 1983).

Chemical and physical properties that help explain observed patterns were introduced in Chapter 2 

and are tabulated in Appendix 2. The properties emphasized are environmental persistence (soil halflife) and mobility in water (represented by the soil organic carbon-water partition coefficient, or Koc). 

The higher the Koc value, the greater the affinity of the compound for soil organic matter, suspended 

particles, and bed sediment—and, thus, a lower tendency to be transported in water. 

•

•

•

•

•

Geographic Distribution

The geographic distribution of each pesticide in streams and ground water is governed 

by the intensity and distribution of its use, 

its chemical and physical properties, and the 

characteristics of the hydrologic system. The 

interactions among these factors are illustrated 

by comparing the findings for several different 

pesticides in relation to their uses and properties. Results for five pairs of the most frequently 

detected pesticides are presented—atrazine and 

metolachlor; simazine and prometon; acetochlor 

and 2,4-D; diazinon and chlorpyrifos; and total 

DDT and dieldrin—representing a wide range of 

use patterns and properties (figs. 4–9 to 4–16). 

These comparative stories provide insights about 

some of the most important pesticides, while also 

illustrating the types and magnitudes of influences that affect all pesticides. 

Occurrence and Distribution in Streams and Ground Water 53
Atrazine and Metolachlor—

The two most heavily used herbicides occurred 

at similar levels in streams, but atrazine was 

more prevalent than metolachlor in ground water, 

probably because of its greater persistence.

Atrazine and metolachlor were the two 

most heavily used herbicides in the United States 

during the 1990s. Most of their agricultural use 

was associated with corn production—about 

85 percent of 75 million lb/yr for atrazine and 

75 percent of 67 million lb/yr for metolachlor 

(fig. 4–3). Both herbicides also have relatively 

low and poorly quantified nonagricultural use—

atrazine is estimated at less than 1 million lb/yr 

(USEPA, 2003a). Uses of metolachlor include 

turf, nurseries, fence rows, and landscaping, and 

uses of atrazine include conifer forestry, Christmas tree farms, sod, golf courses, and residential 

lawns (particularly in the South). Both atrazine 

and metolachlor are highly soluble and mobile in 

water, but atrazine is more persistent than metolachlor, with a soil half-life of 146 days, compared 

with 26 days for metolachlor (Appendix 2). 

Concentrations of both atrazine and metolachlor in agricultural streams closely matched the 

geographic distribution of corn cultivation, where 

applications are greatest (fig. 4–9). Both atrazine 

and metolachlor were also frequently detected in 

urban streams, but at substantially lower concentrations compared with agricultural streams in 

high-use areas, except in parts of the South where 

atrazine is used on turf grasses. Concentrations 

in streams draining watersheds with mixed land 

use most closely resembled those in agricultural 

streams, in large part because many of these 

streams have watersheds with relatively high 

proportions of agricultural land. 

In contrast to their similarity in streams, 

patterns of atrazine and metolachlor were different from each other in ground water (fig. 4–10). 

Metolachlor was detected less frequently than 

atrazine, regardless of land use or depth of 

ground water. This difference probably occurs 

because metolachlor transforms more quickly in 

soil than does atrazine. Metolachlor, therefore, 

is less likely to be transported to ground water, 

although the opposite may be true for some of 

its degradates that appear to be more persistent 

than the parent compound (Kalkhoff and others, 1998). Neither metolachlor nor atrazine was 

detected at the highest frequencies (> 25 percent) 

in ground water underlying large areas of Illinois, 

Indiana, and Ohio, despite their high use in this 

region. This distinct regional pattern, which has 

been noted by several previous studies (Hallberg, 

1989; Burkart and Kolpin, 1993; Baker and others, 1994), is most likely a consequence of the 

widespread use of subsurface drainage systems 

in this area (which move shallow ground water 

rapidly to streams and reduce transport to deeper 

ground water), as well as the presence of lowpermeability glacial till.

Atrazine and metolachlor were heavily used on cropland 

throughout the Corn Belt during the study period.

54 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Occurrence and Distribution in Streams and Ground Water 55

Chapter 4

Atrazine Metolachlor Streams

Agricultural

Urban

Mixed land use

4_09_sw_atrazine_maps

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Stream sites, by watershed land use

Agricultural Urban Mixed 95th percentile concentration, in μg/L

Not detected

< 0.05

0.05 – 0.5

> 0.5

Estimated 1997 agricultural use intensity,

in pounds per square mile per year

"

"

"

"

!

!

!

!

#

#

#

#

EXPLANATION

< 0.09 

0.09 – 4.5

> 4.5 – 45

> 45

Figure 4–9.  Concentrations of both atrazine and metolachlor in agricultural streams closely matched the

geographic distribution of their use on crops. Both atrazine and metolachlor were also often found in urban

streams, but at substantially lower concentrations compared with most agricultural streams. An exception is

atrazine in some urban streams in parts of the South where atrazine was used on turf grasses. Agricultural

use for 1997 was estimated as described in the “Methods” sidebar on p. 53.
56    Pesticides in the Nation’s Streams and Ground Water, 1992–2001

Figure 4–10.  Patterns of atrazine and metolachlor detections were different from each other in ground

water, although both were detected less frequently than expected in the central Corn Belt where the

intensity of use was greatest. Metolachlor was detected less frequently than atrazine, regardless of

land use or depth of ground water, probably because metolachlor is less persistent in soil than atrazine.

Agricultural use for 1997 was estimated as described in the “Methods” sidebar on p. 53.

Atrazine Metolachlor

Agricultural

Urban

4_10_gw_atrazine_maps

Shallow ground-water studies, 

by land use

Agricultural Urban

Major

aquifers Frequency of detection at 0.01 μg/L or 

greater, as percentage of wells in study

Not detected

< 10

10 – 25

> 25

Estimated 1997 agricultural use intensity,

in pounds per square mile per year

"

"

"

"

!

!

!

!

#

#

#

#

EXPLANATION

< 0.09 

0.09 – 4.5

> 4.5 – 45

> 45

Ground Water

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Major aquifers
Chapter 4

Simazine and Prometon—

Although prometon is not registered for 

agricultural use, it frequently occurred in 

agricultural streams and ground water, probably 

because of use for nonagricultural purposes in 

those areas and its high persistence.

Simazine and prometon are commonly 

used herbicides that, compared with atrazine 

and metolachlor, had lower total use and higher 

proportions of nonagricultural use during the 

study period. About 5 million lb/yr of simazine 

were applied for agricultural purposes nationwide (fig. 4–3), compared with about 75 and 

67 million lb/yr of atrazine and metolachlor, 

respectively. Relative to atrazine and metolachlor, simazine is used on a wider variety of 

crops—including corn (about 40 percent of total 

use), citrus orchards (about 35 percent), and 

other orchards and vineyards (about 20 percent). Nonagricultural uses of simazine include 

applications to turf grasses and lawns, roadsides 

and other rights-of-way, and nurseries. Prometon is not registered for agricultural use, but is 

applied for nonagricultural purposes—albeit in 

small amounts—for bare-ground weed control 

around buildings, storage areas and fences, as 

well as along roadways, railroads, and other 

rights-of-way. Both simazine and prometon are 

highly soluble and mobile in water, but prometon is more persistent than simazine, with a soil 

half-life of 932 days, compared with 91 days for 

simazine (Appendix 2). 

The occurrence of simazine in agricultural 

and urban streams was consistent with its geographic patterns of use (fig. 4–11), particularly 

in comparison to the more heavily used atrazine 

(fig. 4–9). For example, concentrations of simazine in agricultural streams in the Corn Belt were 

notably lower than concentrations of atrazine, 

reflecting the lower use of simazine on corn. 

On the other hand, detection frequencies and 

concentrations of simazine in urban streams were 

nearly identical to those of atrazine, reflecting 

generally similar nonagricultural use. Prometon 

was detected less frequently than simazine in 

agricultural streams, at lower concentrations, and 

without the geographic patterns that follow use 

on crops. The prometon detections in agricultural areas probably result from nonagricultural 

applications in these areas. In urban streams, 

prometon was detected at frequencies similar to 

those observed for simazine, atrazine, and diazinon—although at somewhat lower concentrations 

(see figs. 4–11, 4–9, and 4–14, respectively). The 

most likely explanation for the frequent occurrence of prometon is that its high persistence (10 

times that of simazine and more than 5 times that 

of atrazine) results in its prolonged presence in 

watersheds. 

The occurrence and concentrations of simazine in ground water (fig. 4–12) were consistent 

with patterns observed for atrazine and metolachlor (fig. 4–10). Like atrazine and metolachlor, 

detection frequencies were relatively low in 

shallow ground water beneath agricultural areas 

in Illinois, Indiana, and Ohio relative to other 

high use areas—probably because of the common use of subsurface drainage systems and 

widespread presence of glacial till in this region 

(noted earlier). Simazine was generally detected 

more frequently than atrazine and metolachlor 

in Florida and California, which is consistent 

with its higher use in orchards and vineyards in 

those areas. Prometon, consistent with its lack 

of registered agricultural uses, was detected less 

frequently than simazine in shallow ground water 

in agricultural areas. In most urban study areas, 

prometon was detected at similar or greater frequencies than simazine in shallow ground water.

Occurrence and Distribution in Streams and Ground Water 57
58    Pesticides in the Nation’s Streams and Ground Water, 1992–2001

Figure 4–11.  The occurrence and concentrations of simazine in agricultural and urban streams were

consistent with its use, particularly in comparison with the more heavily used atrazine (fig. 4–9.) Prometon

was detected less frequently than simazine in agricultural streams, at lower concentrations, and without the

geographic patterns that follow use on specific crops. Prometon is not registered for agricultural use and no

estimates of agricultural use are shown. Agricultural use of simazine for 1997 was estimated as described in

the “Methods” sidebar on p. 53.

Streams

Agricultural

Urban

Mixed land use

Stream sites, by watershed land use

Agricultural Urban Mixed 95th percentile concentration, in μg/L

Not detected

< 0.05

0.05 – 0.5

> 0.5

Estimated 1997 agricultural use intensity,

in pounds per square mile per year

"

"

"

"

!

!

!

!

#

#

#

#

EXPLANATION

< 0.09 

0.09 – 4.5

> 4.5 – 45

> 45

Simazine Prometon

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4_11_sw_simazine_maps

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!
Occurrence and Distribution in Streams and Ground Water 59

Chapter 4

Figure 4–12.  The occurrence of simazine in ground water was less frequent than atrazine and metolachlor

(fig. 4–10), but the relative patterns were similar, including relatively low detection frequencies in shallow ground

water beneath agricultural areas in Illinois, Indiana, and Ohio, compared with other high-use areas around the

Nation. Prometon was detected less frequently than simazine in shallow ground water within agricultural areas.

Prometon is not registered for agricultural use and no estimates of agricultural use are shown. Agricultural use

of simazine for 1997 was estimated as described in the “Methods” sidebar on p. 53.

4_12_gw_simazine_maps

Simazine Prometon Ground Water

Major aquifers

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Shallow ground-water studies, 

by land use

Agricultural Urban

Major

aquifers Frequency of detection at 0.01 μg/L or 

greater, as percentage of wells in study

Not detected

< 10

10 – 25

> 25

Estimated 1997 agricultural use intensity,

in pounds per square mile per year

"

"

"

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!

!

!

!

#

#

#

#

EXPLANATION

< 0.09 

0.09 – 4.5

> 4.5 – 45

> 45

Agricultural

Urban
2,4-D and Acetochlor—

These two herbicides, which have relatively 

similar chemical and physical properties, have 

different geographic patterns of occurrence in 

streams because of their different use patterns.

The herbicides 2,4-D and acetochlor ranked 

3rd and 5th in national agricultural use during 

the study period (about 41 and 33 million lb/yr, 

respectively, in 1997; fig. 4–3), but their use is 

distributed differently. Acetochlor, which is a relatively new pesticide introduced in 1994, is used 

only on corn, whereas 2,4-D is widely applied 

for multiple agricultural purposes, including 

weed control for pasture (accounting for about 

40 percent of use), wheat (20 percent), corn and 

soybeans (17 percent), as well as other crops and 

fallow land. In addition, 2,4-D has the highest 

documented nonagricultural use of any pesticide 

2,4-D was a commonly 

used herbicide during the 

study period on croplands 

where wheat is grown 

(Photograph copyright by 

Phil Schofield).

(nearly 30 million lb/yr; fig. 4–3). Both 2,4-D 

and acetochlor are relatively soluble and mobile 

in water and neither is particularly persistent, 

with soil half-lives of 7 and 14 days, respectively.

The occurrence and concentrations of these 

two compounds in agricultural and urban streams 

were generally consistent with their patterns 

of use (fig. 4–13). Specifically, relatively high 

concentrations of 2,4-D occurred in agricultural 

streams across the Nation in various high-use 

areas, whereas the highest concentrations of 

acetochlor were generally in the heart of the 

Corn Belt and in other corn-growing areas. Also 

consistent with their use patterns, 2,4-D concentrations were higher in urban streams than 

concentrations of acetochlor. Infrequent low-level 

detections of acetochlor in some urban streams 

may result from relatively minor agricultural use 

within the predominantly urban watersheds or atmospheric 

transport from nearby 

agricultural areas.

Geographic 

results for 2,4-D and 

acetochlor are not 

presented for ground 

water because both 

pesticides were 

detected in less than 

1 percent of the wells 

sampled (fig. 4–2). 

Their infrequent occurrence in ground water 

is probably a result of 

their low persistence. 

For acetochlor, this 

hypothesis is supported 

by the more frequent 

detection of at least 

two of its degradates in 

ground water—relative 

to acetochlor itself—in 

some studies (for 

example, Kalkhoff and 

others, 1998; Groschen 

and others, 2004).

60 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Occurrence and Distribution in Streams and Ground Water 61

Chapter 4

Figure 4–13.  The occurrence and concentrations of 2,4-D and acetochlor in agricultural streams were generally

consistent with their patterns of use. Relatively high concentrations of 2,4-D were observed in agricultural streams

across the Nation, whereas elevated concentrations of acetochlor were generally observed in the heart of the Corn Belt

and in other corn-growing areas. In urban streams, 2,4-D concentrations were generally higher than those of acetochlor.

Agricultural use for 1997 was estimated as described in the “Methods” sidebar on p. 53.

Streams

Agricultural

Urban

Mixed land use

2,4-D Acetochlor

!!

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4_13sw_24D_maps

Stream sites, by watershed land use

Agricultural Urban Mixed 95th percentile concentration, in μg/L

Not detected

< 0.05

0.05 – 0.5

> 0.5

Estimated 1997 agricultural use intensity,

in pounds per square mile per year

"

"

"

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!

!

!

!

#

#

#

#

EXPLANATION

< 0.09 

0.09 – 4.5

> 4.5 – 45

> 45
Chlorpyrifos and Diazinon—

Despite greater use, chlorpyrifos was found 

less frequently than diazinon in water, probably 

because of its greater affinity for particles and 

resulting lower mobility in water.

Chlorpyrifos and diazinon are insecticides 

that were commonly used in both agricultural and 

urban areas during the study period. About 13 

million lb of chlorpyrifos were applied to crops 

in 1997, mostly on corn and cotton (accounting 

for more than 50 percent of national use), with 

the remainder on alfalfa, peanuts, wheat, tobacco, 

and orchards. Less diazinon was used for agriculture (about 1 million lb in 1997), mostly for 

a wide variety of fruits, nuts, and vegetables 

(fig. 4–3). Nonagricultural uses of chlorpyrifos and diazinon totaled about 5 million and 4 

million lb/yr in 2001, respectively (fig. 4–3). 

Both diazinon and chlorpyrifos are substantially 

less mobile in water than the six herbicides just 

discussed, but chlorpyrifos has a greater affinity 

for soil organic matter and particles (higher Koc) 

than diazinon and, thus, a lower solubility and 

mobility in water. Both pesticides have similar 

half-lives in soil—39 days for diazinon and 31 

days for chlorpyrifos (Appendix 2).

The geographic distributions of these insecticides in agricultural and urban streams were 

consistent with their patterns of use (fig. 4–14). 

Of agricultural streams, the highest concentrations of chlorpyrifos were in streams draining the 

corn-growing areas of the central United States 

and the lower Mississippi River Basin, where 

both corn and cotton are grown, and in streams 

draining orchard areas in the West. Concentrations of diazinon in agricultural streams were 

highest in parts of the West where it is intensively used on fruits, nuts, and vegetables. For 

both insecticides, concentrations in most urban 

streams were higher than in most agricultural 

streams, and were similar to those found in agricultural areas with the greatest intensities of use.

In urban streams, diazinon was detected 

about 75 percent of the time, compared with 

about 30 percent for chlorpyrifos (fig. 4–2), even 

though their nonagricultural use was similar. In 

addition, 95th-percentile concentrations equaled 

or exceeded 0.05 μg/L in 23 of 30 urban streams 

for diazinon, compared with only 3 streams for 

chlorpyrifos. In agricultural streams, both chlorpyrifos and diazinon were found at relatively 

similar frequencies and concentrations, despite 

the 10-fold higher use of chlorpyrifos. The markedly greater occurrence of diazinon in proportion to use, compared with chlorpyrifos, may be 

explained by the greater solubility and mobility 

of diazinon in water. Because chlorpyrifos has a 

greater affinity for organic matter than diazinon, 

however, there may have been substantial occurrence and transport of chlorpyrifos in suspended 

sediment in streams that was not observed. As 

discussed in Chapter 3, all NAWQA stream-water 

samples were filtered prior to analysis. 

 Chlorpyrifos and diazinon were rarely 

detected in ground water (less than 1 percent of 

samples; fig. 4–2), so their geographic distributions are not shown. This infrequent occurrence 

is explained by their relatively low persistence 

and low water solubility, as well as their low use 

compared with the major herbicides.

Chlorpyrifos was commonly used on cotton during the 

study period.

Both chlorpyrifos and diazinon were used on apples and 

other orchard crops during the study period.

62 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Occurrence and Distribution in Streams and Ground Water 63

Chapter 4

Figure 4–14.  Concentrations of diazinon, and to a lesser degree chlorpyrifos, in most urban streams were greater than

concentrations in most agricultural streams. Concentrations of diazinon in urban streams were generally similar to those

found in agricultural areas with the greatest intensities of agricultural use. The highest concentrations of diazinon in

agricultural streams were found in the West where it was used on fruits, nuts, and vegetables. The highest concentrations

of chlorpyrifos in agricultural streams were detected in corn-growing areas of the central United States; the lower

Mississippi River Basin, where both corn and cotton are grown; and in streams draining orchard areas in the West.

Agricultural use for 1997 was estimated as described in the “Methods” sidebar on p. 53.

Chlorpyrifos Diazinon

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4_14sw_diazinonchlorpyrifosmaps

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Streams

Agricultural

Urban

Mixed land use

Stream sites, by watershed land use

Agricultural Urban Mixed 95th percentile concentration, in μg/L

Not detected

< 0.05

0.05 – 0.5

> 0.5

Estimated 1997 agricultural use intensity,

in pounds per square mile per year

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EXPLANATION

< 0.09 

0.09 – 4.5

> 4.5 – 45

> 45
DDT and Dieldrin in Bed Sediment—

The geographic distributions of these historically 

used insecticides follow their past agricultural use 

and indicate that use in urban areas probably was 

substantial.

Although the parent pesticides were not 

used in the United States for about 5–20 years 

prior to the beginning of the study period, 

compounds in the DDT group and dieldrin were 

frequently detected in bed sediment. In 1966, 

the combined agricultural use of DDT and DDD 

was about 30 million lb (fig. 4–5), with 66 

percent used on cotton, 9 percent on tobacco, 8 

percent on peanuts, and 17 percent on orchards, 

soybeans, vegetables, potatoes, and other crops. 

The combined agricultural use of dieldrin and 

aldrin (aldrin rapidly transforms to dieldrin in the 

environment) was about 15 million lb in 1966, 

with 92 percent used on corn and 6 percent on 

orchards, vegetables, tobacco, and cotton. Agricultural uses of these insecticides decreased after 

the mid-1960s, and were discontinued by the 

mid-1970s. In addition to their agricultural use, 

aldrin and dieldrin were also widely used for termite control, most intensively in urban areas. Use 

of these compounds as termiticides continued 

until the late 1980s. Although quantitative data 

are not available, DDT also was used extensively 

in nonagricultural applications to control insects 

deemed to be a risk to public health (such as 

mosquitoes), as well as in forestry (U.S. Department of Health and Human Services, 2005; Larson and others, 1997). Compounds in the DDT 

group and dieldrin are all highly persistent—

most with field-dissipation half-lives greater 

than 1,000 days (Nowell and others, 1999)—and 

all have a high affinity for soil organic matter 

(Appendix 2).

Concentrations of total DDT and dieldrin in 

bed-sediment samples from agricultural streams 

correspond reasonably well to both the total 

amounts and the distributions of their historical 

agricultural use (figs. 4–15 and 4–16). Reflecting the higher use of their parent pesticides, 

compounds in the DDT group were detected in 

bed sediment at 49 percent of agricultural stream 

sites, compared with 17 percent for dieldrin. The 

highest total DDT concentrations occurred in 

high-use areas of the Southeast—where cotton, 

tobacco, and peanuts were grown—and in a number of other high-use areas where orchard crops, 

potatoes, vegetables, or specialty crops were 

grown. Dieldrin was found at the highest overall 

concentrations in the Corn Belt, where use of 

aldrin on corn was most intensive.

Although there are few historical data on 

the urban use of organochlorine insecticides, 

the NAWQA bed-sediment results indicate that 

it probably was substantial. Compounds in the 

DDT group and dieldrin were found at higher frequencies and generally higher concentrations in 

urban streams than in agricultural streams, with 

the exception of DDT in some streams draining 

agricultural watersheds that had high DDT use 

in the past. Compounds in the DDT group were 

detected in 72 percent of samples from urban 

streams, compared with 42 percent for dieldrin.

For most streams with mixed land use in 

their watersheds, the concentrations of total DDT 

and dieldrin were generally similar to those in 

agricultural streams, but lower than those in 

urban streams. Streams in undeveloped watersheds had the lowest concentrations of these 

compounds.

In addition to agricultural uses, DDT also was 

applied to wetlands and marshes to control 

mosquitoes (photograph courtesy of the Tennessee 

Valley Authority Historic Collection, 1938).

64 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Occurrence and Distribution in Streams and Ground Water 65

Chapter 4

Figure 4–15.  Total DDT concentrations in bed sediment were generally higher in urban streams than in

agricultural and mixed-land-use streams, with the exception of a few streams draining watersheds in areas

that had high agricultural use of DDT plus DDD in the past. The distribution of concentrations of total DDT

found in bed sediment of agricultural streams corresponded reasonably well to both the total amount and

the distribution of historical agricultural use of DDT plus DDD. Total DDT concentrations were highest in

high-use areas of the Southeast where cotton, tobacco, and peanuts were grown, and in a number of other

high-use areas where orchard crops, potatoes, vegetables, or specialty crops were grown. Historical use

for the late 1960s was estimated as described in the “Methods” sidebar on p. 53.

Agriculture Agriculture

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Total DDT in bed sediment

Agricultural Mixed land use

4_15_DDT

Urban Undeveloped

Estimated historical agricultural 

use intensity, in pounds per square 

mile per year

EXPLANATION

Stream sites, by watershed land use Concentration, in μg/kg

dry weight

Not detected or < 1

1 – 5

> 5

< 0.09 

0.09 – 4.5

> 4.5 – 45

> 45
66    Pesticides in the Nation’s Streams and Ground Water, 1992–2001

Dieldrin in bed sediment

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4_16_Dieldrinbed_sediment_maps

Agriculture Agricultural Mixed land use Agriculture

Urban Undeveloped

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Agricultural Urban Mixed Undeveloped

"

"

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Estimated historical agricultural 

use intensity, in pounds per square 

mile per year

EXPLANATION

Stream sites, by watershed land use Concentration, in μg/kg 

dry weight

Not detected or < 1

1 – 5

> 5

< 0.09 

0.09 – 4.5

> 4.5 – 45

> 45

Figure 4–16.  Dieldrin in bed sediment generally occurred at higher frequencies and higher concentrations

in urban streams than in most agricultural streams. Concentrations of dieldrin found in bed sediment of

agricultural streams corresponded reasonably well to the distribution of historical agricultural use of aldrin

plus dieldrin. In agricultural streams, dieldrin was detected most frequently and at the highest overall

concentrations in the Corn Belt, where past use of aldrin on corn was most intensive. Historical use for the

late 1960s was estimated as described in the “Methods” sidebar on p. 53.
Chapter 5

The occurrence of pesticides in streams and ground water, 

which was characterized in relation to land use and the geo-

graphic patterns in pesticide use in Chapter 4, is further compli-

cated by three additional factors: strong seasonal patterns, the 

prevalence of mixtures of pesticides, and the frequent occur-

rence of degradates. Seasonal patterns occur year after year 

in most streams and dictate the timing of the highest pesticide 

concentrations; mixtures of multiple pesticide compounds 

are found more often than individual pesticides; and pesticide 

degradates may occur more frequently and at higher concentra-

tions than their parent compounds, particularly in ground water. 

These complexities need to be understood and considered when 

assessing the potential effects of pesticides on water quality.

This chapter provides an 

overview of national findings and 

selected case studies regarding 

seasonal patterns, mixtures, and 

degradates of pesticides.

5

Complexities: Seasonality, Mixtures, 

and Degradates

Photograph © Kemer Nelson, Aequalis Aerial Photography, Tacoma, Washington.
Figure 5–1. The timing and magnitude of seasonal pulses in the concentrations of herbicides and insecticides 

differed between agricultural and urban streams. Herbicide concentrations tended to be higher and seasonal 

patterns more pronounced in agricultural streams, but insecticide concentrations generally were higher in urban 

streams. Median and 75th percentile concentrations were determined after aggregating the total concentrations 

of herbicides and insecticides for all samples from agricultural streams and for all samples from urban streams.

0

1

2

3

Concentration,

in micrograms per liter

0

1

2

3

Concentration,

in micrograms per liter

Concentration,

in micrograms per liter

Concentration,

in micrograms per liter

Agricultural streams

9.3 6.8

Urban streams

0.1

0

0.2

0.3

0.1

0

0.2

0.3

3.3

January

February

March

April

May

June

July

August

September

October

NovemberDecember

January

February

March

April

May

June

July

August

September

October

NovemberDecember

January

February

March

April

May

June

July

August

September

October

NovemberDecember

January

February

March

April

May

June

July

August

September

October

NovemberDecember

75th percentile

Median

75th percentile

Median

Total Herbicides

National overview of seasonal patterns in streams

Agricultural streams Urban streams

Total Insecticides

Seasonal Patterns in Streams

Concentrations of pesticides in streams typically follow marked seasonal patterns year after 

year. These patterns generally are characterized 

by long periods of low or undetectable concentrations, punctuated by a few weeks or months of 

higher concentrations—a seasonal pulse. Such 

patterns are governed primarily by the timing and 

intensity of pesticide use in relation to hydrologic factors that affect the transport of pesticides 

to streams. Key hydrologic factors include the 

timing and amount of runoff from rainfall and 

irrigation, the presence or absence of surface or 

subsurface drainage systems, and the degree of 

interaction between streams and ground water. 

Seasonal patterns are important to characterize 

because they dictate the timing and duration of 

the highest concentrations of pesticides that may 

affect the suitability of water for humans, aquatic 

life, and wildlife. 

NAWQA findings show that concentrations 

of pesticides in agricultural and urban streams 

across the Nation usually were highest during 

the growing season and lowest during the winter 

(fig. 5–1). The highest concentrations of herbicides—generally higher in agricultural streams 

than in urban streams—usually occurred during 

68 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 5

Figure 5–2. Atrazine and its degradates dominated herbicide concentrations in Bogue Chitto Creek, an agricultural stream 

(Mobile River Basin), with concentrations peaking in the spring following applications on corn fields and gradually declining 

throughout the summer and winter. In nearby Cahaba Valley Creek, an urban stream, herbicide concentrations were highest 

during November–April, and the dominant herbicide was simazine (Atkins and others, 2004).

Bogue Chitto Creek, Alabama (agricultural)

Seasonal patterns of herbicides in an agricultural and an urban stream

Cahaba Valley Creek, Alabama (urban)

March

April

May

June

July

August

September

October

NovemberDecember

January

February

March

April

May

June

July

August

September

October

NovemberDecember

January

February

14

12

10

8

6

4

2

0

2.0

1.5

1.0

0.5

0

Median monthly concentration, 

in micrograms per liter

Median monthly concentration, 

in micrograms per liter

Others

Cyanazine

Atrazine and degradates

Others

2,4-D

Atrazine and degradates

Simazine

April–July. In contrast, the highest concentrations of insecticides—generally higher in urban 

streams than in most agricultural streams—usually occurred at various times over a longer 

period, from March through September. Differences that may occur in seasonal patterns 

between agricultural and urban streams, even 

within the same geographic area, are illustrated by findings from the Mobile River Basin 

(fig. 5–2). Numerous additional examples of 

seasonal patterns have been characterized for 

streams in different parts of the country in individual NAWQA studies (see reports for the 51 

NAWQA Study Units: http://water.usgs.gov/

nawqa/nawqasum/).

Collecting a water sample 

from Cahaba Valley Creek.

Complexities: Seasonality, Mixtures, and Degradates 69
Geographic Variability of Seasonal 

Patterns 

Although the occurrence and concentrations of pesticides followed distinct seasonal 

patterns in most of the agricultural and urban 

streams sampled, the specific timing and magnitude of the observed patterns varied regionally 

and locally. This variability results from differences in such factors as the timing and amounts 

of pesticide use, climate, and the frequency and 

magnitude of runoff from rainstorms or irrigation. Seasonal patterns were particularly consistent within regions in which climate, land use, 

and crop types are relatively uniform, such as 

in the Corn Belt. The accompanying map and 

graphs (fig. 5–3) show examples that illustrate 

regional consistency, variability among streams, 

and land-use influences on seasonality using 

findings for atrazine, prometon, and diazinon in 

selected streams. 

Corn Belt Streams—Concentrations of atrazine, the dominant herbicide used in the Corn 

Belt during the study period, typically peaked 

after applications in the spring, as shown in 

four streams draining parts of Iowa, Indiana, 

Ohio, and Mississippi. Atrazine use generally is consistent from year to year, closely 

following annual patterns of weather and weed 

growth. Concentrations of prometon in these 

streams had weaker seasonal patterns and were 

lower than concentrations of atrazine, because 

prometon is used in smaller quantities and for 

a variety of nonagricultural purposes. Concentrations of diazinon were low or undetectable during most or all of the year in all Corn 

Belt streams except the Maumee River. The 

Maumee River has more urban land in its 

watershed compared with the other Corn Belt 

streams shown and was probably influenced 

more by nonagricultural diazinon use than the 

other streams.

•

Urban Streams—Seasonal concentration patterns in three streams draining urban areas 

in Virginia, Georgia, and Nevada were more 

erratic than those observed in most agricultural areas. This was typical of the urban 

streams sampled nationwide, probably because 

pesticides are applied more sporadically in 

residential and commercial settings to control outbreaks of insects and weeds as they 

occur. Prometon and diazinon were generally 

detected at higher concentrations than atrazine 

in Las Vegas Wash and Accotink Creek, with 

the highest concentrations occurring during 

spring or summer. Sope Creek had higher 

concentrations of atrazine and lower concentrations of prometon than the other urban 

streams. The higher atrazine concentrations in 

Sope Creek may have resulted from the use of 

atrazine for treating turf grass in Georgia. 

Palouse River—Concentrations of prometon 

and atrazine were highest during winter and 

spring in the Palouse River in Washington, but 

were low overall compared with agricultural 

streams draining the Corn Belt. The Palouse 

River drains mostly nonirrigated cropland 

where wheat and other grains are the primary 

crops and pesticide use is relatively low.

Orestimba Creek—Diazinon concentrations 

peaked during early winter and midsummer 

in Orestimba Creek, California, which drains 

irrigated farmlands dominated by orchards, 

vegetables, and alfalfa. Diazinon was used 

extensively in this watershed on almond 

orchards in January and February and on vegetable crops during the summer.

•

•

•

70 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 5

Figure 5–3. Examples of seasonal patterns in concentrations of atrazine, prometon, and diazinon in selected agricultural and 

urban streams illustrate (1) the regional consistency of patterns for atrazine within the Corn Belt, an area that has relatively 

uniform agricultural practices; (2) the variability among streams in different regions of the country; and (3) the differing 

influences of land use on seasonality. Median concentrations for each month were computed from at least 6 years of data for 

each site and nondetections are plotted at 0.001 μg/L.

Las Vegas

Wash, NV

Orestimba Creek, CA

Yazoo River, MS 10

January

March

May

September

November

July

1

0.1

0.01

0.001

10

1

0.1

0.01

0.001

10

1

0.1

0.01

0.001

Palouse River, WA

Median monthly concentration,

in micrograms per liter

Iowa River, IA

Pesticides

 Atrazine

 Prometon

 Diazinon

Stream site

 Agricultural

 Urban

Sugar Creek, IN Maumee River, OH

Sope Creek, GA

Accotink Creek, VA

10

1

0.1

0.01

0.001

10

1

0.1

0.01

0.001

10

1

0.1

0.01

0.001

10

1

0.1

0.01

0.001

10

1

0.1

0.01

0.001

10

1

0.1

0.01

0.001

J M M J S N

J M M J S N

J M M J S N J M M J S N

J M M J S N

J M M J S N

J M M J S N J M M J S N J M M J S N

Seasonal patterns of atrazine, prometon, and diazinon in selected streams

5_03

EXPLANATION

Complexities: Seasonality, Mixtures, and Degradates 71
100

1,000

10,000

100,000

1,000,000

0.001

1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

0.01

0.1

1

10

100

Concentration, in micrograms per liter

Streamflow, in cubic feet per second

Seasonal patterns in atrazine and chlorpyrifos in the White River, Indiana

Atrazine

Chlorpyrifos

Streamflow

Chlorpyrifos not detected

Figure 5–4. Atrazine concentrations in the White River (White River Basin) followed the same pattern each year 

during 1992–2001, corresponding to its use for weed control on nearly all of the corn acreage in the watershed 

each spring. In contrast, seasonal patterns in concentrations of chlorpyrifos and other insecticides tend to be more 

variable because insecticides are typically applied more sporadically than herbicides.

Repetition of Seasonal Patterns

Seasonal patterns of pesticide concentrations in each particular stream generally repeat 

with varying degrees of consistency each year, as 

long as the pesticides are still in use. For example, atrazine concentrations in the White River 

followed the same pattern each year from 1992 

to 2001 (fig. 5–4). Corn is planted in the region 

between mid-April and the end of May, and atrazine is applied each year to nearly all of the corn 

acreage during this time period. Runoff resulting 

from rainfall in May and June transports atrazine 

to streams, giving rise to the highest concentrations of the year during and after application. In 

contrast, patterns in chlorpyrifos concentrations 

in the White River were less regular (fig. 5–4) 

because the insecticide is generally applied only 

if and when it is needed to control outbreaks of 

corn root worm. 

Importance of Seasonal Patterns

Seasonal patterns in pesticide concentrations 

are important to understand because they may 

affect the management of water quality for some 

drinking-water supplies and often define critical 

conditions of pesticide exposure for aquatic life 

in a stream. Although NAWQA did not measure pesticide concentrations at drinking-water 

intakes, NAWQA results for the wide range of 

streams sampled indicate that seasonal pulses 

of pesticide concentrations probably occur in 

some streams that are used as sources of drinking 

water—primarily those with substantial agricultural or urban land use in their watersheds. For 

drinking-water sources where seasonal patterns 

are evident, seasonal monitoring is important to 

support water-quality management decisions. 

For example, some drinking-water utilities that 

withdraw water from streams in agricultural areas 

employ specific management strategies to avoid 

use of stream water, or to increase treatment of 

the water, during known seasonal periods of high 

concentrations in source waters. 

The seasonal timing of elevated pesticide concentrations in relation to the timing of 

changes in populations and life stages of aquatic 

organisms may largely determine whether pesticides have a substantial effect on aquatic life in 

a stream. USEPA’s Office of Pesticide Programs, 

for example, evaluates potential acute effects of 

exposure on the basis of peak concentration, and 

potential chronic effects on the basis of the peak 

21-day average for invertebrates and the peak 60-

day average for fish (see Chapter 6). As indicated 

by figures 5–1 through 5–4, in most streams, 

these daily and multiday average concentrations 

are most likely to be approached or exceeded 

during relatively distinct seasonal periods for 

each pesticide. Knowledge of the seasonal timing 

of the highest concentrations for each pesticide—

together with an understanding of the life stages 

of aquatic organisms present in each season—can 

be used to target and refine assessments of 

potential effects, and to design efficient pesticide 

monitoring strategies that will yield reliable estimates of the concentration statistics required for 

site-specific risk assessments. 

72 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 5

Mixtures of Pesticides

Assessment of the effects of pesticides 

on water quality is further complicated by the 

simultaneous occurrence of multiple pesticides 

and degradates as mixtures. The mixtures result 

from the use of different pesticides for multiple 

purposes within a watershed or ground-water 

recharge area. Pesticides generally occur more 

often as mixtures than as individual compounds. 

As a result, evaluation of the potential effects of 

mixtures of pesticides and other contaminants is 

an increasingly important component of the risk 

assessment methods used by USEPA, the Agency 

for Toxic Substances and Disease Registry 

(ATSDR), and other agencies (see Chapter 6). 

Consistent with the results for individual 

compounds discussed in Chapter 4, mixtures of 

pesticides were detected more often in streams 

than in ground water (fig. 5–5) and at relatively 

similar frequencies in streams draining areas of 

agricultural, urban, and mixed land use. More 

than 90 percent of the time, water from streams 

in these developed land-use settings had detections of 2 or more pesticides or degradates; about 

70 percent of the time, streams had 5 or more, 

and about 20 percent of the time, streams had 

detections of 10 or more pesticides or degradates. Mixtures also were found in streams 

draining undeveloped watersheds, but with far 

fewer compounds—about 25 percent of the time, 

undeveloped streams had detections of 5 or more 

pesticides or degradates, and no samples had 

more than 10.

In ground water, pesticide mixtures were 

detected most frequently in shallow wells in 

agricultural and urban areas—47 percent of wells 

sampled in agricultural areas and 37 percent of 

wells in urban areas had detections of 2 or more 

pesticides or degradates. Only about 9 percent 

of the wells sampled in these areas contained 5 

or more pesticides or degradates, and less than 1 

percent contained more than 10. Consistent with 

the slow rate of ground-water movement and the 

resulting greater opportunities for sorption and 

transformation with increasing residence time, 

co-occurrences of multiple pesticides and degradates were less frequent in wells that tap major 

aquifers—only about 20 percent of such wells 

had detections of 2 or more pesticides or degradates. Mixtures were least prevalent in wells 

sampled in undeveloped areas.

Mixtures of organochlorine pesticide compounds also were common in fish-tissue samples 

from most streams (fig. 5–6). About 90 percent 

of fish samples collected from urban streams 

contained 2 or more pesticide compounds and 

33 percent contained 10 or more. Similarly, 75 

percent of fish samples from streams draining 

watersheds with agricultural and mixed land use 

contained 2 or more pesticide compounds and 10 

percent had 10 or more. As with water samples, 

mixtures were detected least often in fish from 

undeveloped streams, in which 2 or more compounds were detected in about 25 percent of the 

fish-tissue samples. 

The potential for effects of mixtures on 

humans, aquatic life, and fish-eating wildlife is 

ultimately determined by the specific combinations of compounds that occur together, their 

concentrations, and when and where they occur. 

A unique mixture is defined in this report as a 

combination of 2 or more particular compounds 

detected in a given sample, regardless of whether 

other compounds were also detected in the same 

sample (Squillace and others, 2002). For example, a sample containing compounds A, B, and 

Figure 5–5. Mixtures of pesticide compounds analyzed 

in water were common in streams draining watersheds 

with agricultural, urban, and mixed land use. More 

than 90 percent of the time, water from streams in 

these land-use settings had detections of 2 or more 

pesticides or degradates, and almost 20 percent of the 

time, streams had detections of 10 or more. Mixtures 

were less common in ground water, but shallow wells 

in agricultural and urban areas had the most frequent 

occurrences of mixtures among all ground-water 

samples. 

Frequency of mixtures in water

Number of pesticide compounds in sample

1 2 3 4 5 6 7 8 9 10

0

25

50

75

100

Percentage of time (streams) or 

samples (ground water) with 

equal or greater number of detectable

pesticide compounds

Land use Streams Ground

water

Agricultural

Urban

Undeveloped

Mixed land use or

major aquifers

Complexities: Seasonality, Mixtures, and Degradates 73
C contains four unique mixtures—AB, AC, BC, 

and ABC. The number of unique mixtures in one 

sample can be very large if many compounds are 

detected. A sample with 2 compounds has only 

one mixture, but a sample with five compounds 

contains 26 unique mixtures. Unique mixtures 

examined in this report were limited to those 

composed of the most commonly detected pesticide compounds—specifically, 25 compounds in 

water and 15 in fish tissue (figs. 4–2 and 4–4)—

and were further limited to unique mixtures that 

occurred at least 2 percent of the time in streams 

or in at least 2 percent of samples for ground 

water and fish tissue.

pesticides detected in shallow ground water 

within agricultural areas.

 The number of unique mixtures varied 

with land use. For example, the greatest number of unique mixtures occurred in agricultural 

streams, probably because of the wide variety of 

agricultural settings represented, each involving 

the use of different combinations of pesticides 

(note the logarithmic scale in fig. 5–7). In ground 

water, however, the greatest number of unique 

mixtures occurred in shallow wells within urban 

areas, resulting primarily from the detection of 

urban herbicides that were not used or detected 

as frequently in agricultural settings. Major 

aquifers had the fewest mixtures, consistent with 

the lower frequencies of detection for individual 

compounds in these deeper ground waters.

The number of unique mixtures that can be 

detected is strongly influenced by the detection 

level for individual pesticides. In ground water, 

where pesticide concentrations usually are low, 

only 1 unique mixture of 2 compounds (atrazine 

and deethylatrazine in agricultural areas) was 

identified when the analysis was restricted to 

detected concentrations greater than 0.1 μg/L. In 

streams, however, many unique mixtures were 

detected—even when evaluating only detections greater than 0.1 μg/L (fig. 5–7). At the 

0.1 μg/L detection level, greater distinctions were 

evident between land-use settings. For example, 

about 50 unique 5-compound mixtures were 

detected in agricultural streams when only individual pesticides at concentrations greater than 

0.1 μg/L are considered (compared with more 

than 6,000 when including all detections at any 

concentration). In urban streams, only 6 unique 

5-compound mixtures were detected above the 

0.1 μg/L level.

The most frequent contributors to mixtures, 

not surprisingly, are the individual pesticides 

that were detected most often (fig. 5–8). These 

include the herbicides atrazine (and its degradate deethylatrazine), metolachlor, simazine, 

and prometon, each of which was present in 

more than 30 percent of all mixtures found in 

agricultural and urban areas, and in both streams 

and ground water. Also present in more than 30 

percent of the mixtures were cyanazine, alachlor, 

metribuzin, and trifluralin in agricultural streams, 

and dacthal and the insecticides diazinon, chlorpyrifos, carbaryl, and malathion in urban streams. 

The most notable difference between urban 

and agricultural streams was the more common 

occurrence of insecticides in mixtures found in 

urban streams—consistent with the generally 

more frequent occurrence of insecticides in urban 

Figure 5–6. Mixtures of organochlorine pesticide 

compounds were detected in samples of whole 

fish from most streams. Fish in urban streams 

had the greatest numbers of organochlorine 

compounds detected, with about 90 percent of 

fish samples containing 2 or more compounds 

and 33 percent containing 10 or more. Streams in 

areas of agricultural and mixed land use had about 

75 percent of fish-tissue samples with 2 or more 

compounds and 10 percent with 10 or more. Percentage of samples with equal or greater number of detectable pesticide compounds

0

25

50

75

100

Number of pesticide compounds in sample

1 2 3 4 5 6 7 8 9 10

Land use Streams

5_06abmixBSOV

Agricultural

Urban

Undeveloped

Mixed

Frequency of mixtures in fish tissue

Unique Mixtures in Water

Streams have more unique mixtures than 

ground water (fig. 5–7), which is consistent with 

the more frequent detection of pesticides and 

their degradates in stream water compared with 

ground water. Analysis of all detections indicates 

that more than 6,000 unique mixtures of five 

pesticides were detected in agricultural streams, 

compared with only one unique mixture of five 

74 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 5

0 1 2 3 4 5 6 7 8 9 10

Number of pesticide compounds in mixture

Number of unique mixtures

1

10

100

1,000

10,000 Streams

All

detections

Land use Detections

 0.1 μg/L

Agricultural

Urban

Mixed land use or

major aquifers

1

0 1 2 3 4 5 6 7 8 9 10

10

100

Number of pesticide compounds in mixture

Number of unique mixtures

Ground Water

0bSTGW

Occurrence of unique mixtures in water

 0.1 μg/L

Figure 5–7. The number of unique mixtures detected in water was much greater in streams than in ground water. Considering all 

detections, more than 6,000 unique 5-compound mixtures were found in samples from agricultural streams, whereas only 1 unique 

5-compound mixture was detected in shallow ground water beneath agricultural areas. Considering only pesticides detected at 

concentrations greater than or equal to 0.1 μg/L, the number of unique mixtures was far less. These graphs include only those unique 

mixtures that were composed of the 25 most prevalent pesticides and were detected in at least 2 percent of the samples.

streams. A notable difference in ground water 

between urban and agricultural areas was the 

occurrence of tebuthiuron, which was present 

in about 35 percent of the mixtures detected in 

wells in urban areas, but in less than 2 percent of 

the mixtures in wells in agricultural areas.

 The unique mixtures detected most frequently in streams and ground water are summarized in table 5–1. This assessment is limited to 

the pesticides measured by NAWQA and by the 

sensitivity of the analytical method for each pesticide. For example, the analysis under-represents 

the contributions of 2,4-D, bentazon, bromacil, 

carbaryl, diuron, and norflurazon to mixtures, 

relative to the other pesticides, because these 

compounds were only detectable at higher concentrations. Most notably, 2,4-D was one of the 

most prevalent components of mixtures in both 

agricultural and urban streams at concentrations 

of 0.1 μg/L or greater. This implies that 2,4-D 

is also likely to be one of the most important 

contributors to mixtures at lower concentrations 

as well, but the low concentrations could not be 

measured.

Combinations of agricultural and urban land uses, such 

as those on Oahu, Hawaii, result in use of many different 

pesticides, leading to complex mixtures of pesticide 

compounds in streams and ground water (Landsat 

satellite image from the Pacific Disaster Center).

Complexities: Seasonality, Mixtures, and Degradates 75
Figure 5–8. The most common components of mixtures, not surprisingly, were the pesticides and degradates that were detected most 

often. The most frequent contributors to unique mixtures were the herbicides atrazine (and deethylatrazine), metolachlor, simazine, 

and prometon—all of which were detected in more than 30 percent of all unique mixtures found in agricultural and urban areas and 

in streams and ground water. The most notable differences between agricultural and urban areas were (1) the greater contribution of 

insecticides to the mixtures detected in urban streams, and (2) the greater contribution of tebuthiuron to the mixtures found in shallow 

ground water in urban areas. This analysis is based on detections at any concentration, but includes only those unique mixtures that 

were composed of the 25 most prevalent pesticides and were detected in at least 2 percent of samples.

Atrazine

Deethylatrazine

Metolachlor

Cyanazine

Alachlor

Acetochlor

Metribuzin

Bentazon

EPTC

Trifluralin

Molinate

Norflurazon

Simazine

Prometon

Tebuthiuron

2,4-D

Diuron

Dacthal

Bromacil

Diazinon

Chlorpyrifos

Carbofuran

Carbaryl

Malathion

Dieldrin

Streams

Composition of unique mixtures in water

Ground Water

Agricultural Urban Agricultural Urban

Contribution to unique mixtures, as a percentage of unique mixtures

Agricultural

herbicides

Urban

herbicides Insecticides

5_08mainpest

0 25 50 75 100 0 25 50 75 100 0 25 50 75 100 0 25 50 75 100

76 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 5

Table 5–1. The most common unique mixtures of pesticides and degradates found in stream water and 

ground water illustrate the diversity and complexity of mixtures that occur in agricultural and urban areas. 

The mixtures detected most frequently for each number of compounds are shown for each land use, 

with all detections included, regardless of concentration. These most common unique mixtures serve as 

examples, rather than as a comprehensive compilation of all the most important mixtures, because other 

mixtures occurred almost as frequently. 

Mixture

Frequency of detection

(percentage of time for streams, or samples for 

ground water)

Streams Ground water

Urban Agricultural Urban Agricultural

2-compound mixtures

Atrazine Prometon 79 50 15 10

Prometon Simazine 75 41 10 7

Atrazine Simazine 74 64 17 18

Atrazine Metolachlor 55 77 8 15

Atrazine Deethylatrazine 53 77 26 39

Deethylatrazine Simazine 49 57 15 17

Deethylatrazine Metolachlor 42 69 7 14

3-compound mixtures

Atrazine Prometon Simazine 68 41 9 7

Atrazine Diazinon Prometon 60 10 1 0

Atrazine Diazinon Simazine 59 16 1 0

Diazinon Prometon Simazine 59 9 1 0

Atrazine Deethylatrazine Prometon 50 48 12 9

Atrazine Deethylatrazine Simazine 48 57 15 16

Atrazine Metolachlor Simazine 48 57 6 7

Atrazine Deethylatrazine Metolachlor 41 69 7 14

4-compound mixtures

Atrazine Diazinon Prometon Simazine 53 9 1 0

Atrazine Deethylatrazine Prometon Simazine 46 39 8 7

Atrazine Metolachlor Prometon Simazine 43 38 4 4

Atrazine Deethylatrazine Metolachlor Prometon 39 45 5 6

Atrazine Deethylatrazine Metolachlor Simazine 37 52 5 7

Alachlor Atrazine Deethylatrazine Metolachlor 14 42 0 2

5-compound mixtures

Atrazine Carbaryl Diazinon Prometon Simazine 36 2 0 0

Atrazine Deethylatrazine Diazinon Prometon Simazine 35 8 1 0

Atrazine Deethylatrazine Metolachlor Prometon Simazine 35 37 4 4

Atrazine Diazinon Metolachlor Prometon Simazine 35 8 0 0

Atrazine Deethylatrazine Prometon Simazine Tebuthiuron 28 16 2 1

Atrazine Deethylatrazine Metolachlor Simazine Tebuthiuron 22 19 2 1

Alachlor Atrazine Deethylatrazine Metolachlor Prometon 13 33 0 1

Alachlor Atrazine Deethylatrazine Metolachlor Simazine 13 33 0 1

Alachlor Atrazine Deethylatrazine Prometon Simazine 12 26 0 1

Atrazine Cyanazine Deethylatrazine Metolachlor Simazine 5 33 1 1

Complexities: Seasonality, Mixtures, and Degradates 77
Unique Mixtures in Fish Tissue

The numbers of unique mixtures of organochlorine pesticide compounds found in fish tissue 

are summarized in figure 5–9. Each individual 

parent compound, degradate, and by-product 

included in figure 4–4 was counted separately. 

Urban streams had more unique mixtures of these 

compounds in fish than streams draining areas 

with agricultural or mixed land use. For example, 

about 1,400 unique 5-compound mixtures were 

found in fish from urban streams, whereas 

streams in areas with agricultural or mixed land 

use had fewer than 800 unique 5-compound 

mixtures.

The relative contributions of most organochlorine compounds to mixtures in fish were 

about the same for urban and agricultural streams 

Figure 5–9. The numbers of unique mixtures of 

organochlorine pesticide compounds found in 

whole-fish tissue samples were greater in urban 

streams than in streams with agricultural or 

mixed-land-use watersheds. For example, about 

1,400 unique 5-compound mixtures were found in 

fish from urban streams, whereas fewer than 800 

unique 5-compound mixtures were detected in 

fish from agricultural and mixed-land-use streams. 

This analysis includes all detections, but only 

those unique mixtures that were composed of the 

15 organochlorine pesticide compounds included 

in figure 4–4 and that were detected in at least 

2 percent of samples. 5_09_mix2pct_fish

0

250

500

750

1,000

1,250

1,500

Land use All detections

Agricultural

Urban

Mixed

0 1 2 3 4 5 6 7 8 9 10

Number of pesticide compounds in mixture

Number of unique mixtures

Occurrence of unique mixtures in 

fish tissue

Figure 5–10. For the most part, compounds that 

contributed to unique mixtures found in fish tissue 

were similar for agricultural and urban streams, 

and 10 of the 15 compounds included in the analysis 

contributed to more than 30 percent of the mixtures 

in both land-use settings. 

p,p’-DDE

p,p’-DDD

p,p’-DDT

o,p’-DDE

o,p’-DDD

o,p’-DDT

cis-Chlordane

trans-Chlordane

cis-Nonachlor

trans-Nonachlor

Oxychlordane

Dieldrin

Heptachlor epoxide

Pentachloroanisole

Hexachlorobenzene

Agricultural Urban

Contribution to unique mixtures, 

as a percentage of mixtures

Total DDT

Total

Chlordane

Composition of unique mixtures in 

fish tissue

0 25 50 75 100 0 25 50 75 100

(fig. 5–10). This reflects the fact that most 

of these compounds originated with a small 

number of pesticide products that were applied 

many years ago in both land-use settings. The 

most frequent contributors were compounds 

derived from formulations of DDT and DDD 

(especially the p,p′ isomers), dieldrin (resulting 

from use of either aldrin or dieldrin), chlordane, 

and heptachlor epoxide (resulting from use of 

heptachlor)(table 5–2). The most notable difference between agricultural and urban streams 

was the greater importance in urban streams of 

hexachlorobenzene (an industrial compound, as 

well as a fungicide) and both o,p′- and p,p′-DDT. 

The greater prevalence of DDT isomers in 

mixtures in urban streams, relative to agricultural 

streams, is consistent with the finding that the 

parent compounds (o,p′- and p,p′-DDT) made 

up a higher proportion of total DDT residues in 

fish from urban streams (16 percent) than in fish 

from streams with agricultural, undeveloped, or 

mixed-land-use watersheds (2–3 percent). These 

results indicate either that DDT was applied more 

recently in urban watersheds, or that there has 

been more recent transport of DDT-contaminated 

soils to the streams in urban areas than to streams 

in most agricultural watersheds. (DDT breaks 

down more rapidly in the absence of dissolved 

oxygen and is, in general, less persistent in 

aquatic sediment than in soil.) 

78 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 5

Table 5–2. The most common unique mixtures of organochlorine pesticide compounds found in fish 

tissue were dominated by compounds derived from commercial formulations of DDT, DDD, and chlordane, 

as well as dieldrin. The mixtures detected most frequently for each number of compounds are shown for 

each land use. These most common mixtures serve as examples, rather than having unique importance, 

because many other mixtures occur almost as frequently.

Mixture

Frequency of detection

 (percentage of samples) 

Fish tissue

Urban streams Agricultural 

streams

2-compound mixtures

p,p'-DDE trans-Nonachlor 72 49

cis-Chlordane p,p'-DDE 72 37

cis-Chlordane trans-Nonachlor 72 35

p,p'-DDD p,p'-DDE 64 59

Dieldrin p,p'-DDE 53 59

3-compound mixtures

cis-Chlordane p,p'-DDE trans-Nonachlor 68 35

cis-Chlordane p,p'-DDD p,p'-DDE 60 28

cis-Chlordane p,p'-DDD trans-Nonachlor 57 27

p,p'-DDD p,p'-DDE trans-Nonachlor 57 37

Dieldrin p,p'-DDE trans-Nonachlor 53 42

Dieldrin p,p'-DDD p,p'-DDE 40 47

4-compound mixtures

cis-Chlordane p,p'-DDD p,p'-DDE trans-Nonachlor 56 27

cis-Chlordane p,p'-DDE trans-Chlordane trans-Nonachlor 51 25

cis-Chlordane p,p'-DDE p,p'-DDT trans-Nonachlor 51 9

cis-Chlordane Dieldrin p,p'-DDE trans-Nonachlor 50 33

cis-Chlordane Dieldrin p,p'-DDD p,p'-DDE 40 27

Dieldrin p,p'-DDD p,p'-DDE trans-Nonachlor 40 34

5-compound mixtures

cis-Chlordane p,p'-DDD p,p'-DDE trans-Chlordane trans-Nonachlor 46 19

cis-Chlordane p,p'-DDD p,p'-DDE p,p'-DDT trans-Nonachlor 44 9

cis-Chlordane Dieldrin p,p'-DDE trans-Chlordane trans-Nonachlor 43 24

cis-Chlordane Dieldrin p,p'-DDD p,p'-DDE trans-Nonachlor 40 26

cis-Chlordane cis-Nonachlor Dieldrin p,p'-DDE trans-Nonachlor 39 23

Complexities: Seasonality, Mixtures, and Degradates 79
Importance of Mixtures

The frequent detection of pesticide mixtures 

in NAWQA samples from streams and ground 

water indicates that aquatic life, fish-eating 

wildlife, and potentially humans, are exposed 

primarily to mixtures of pesticides, rather than to 

individual compounds. As examined in Chapter 

6, determining the potential effects of mixtures is 

an increasingly important aspect of risk-assessment procedures for pesticides. These procedures 

generally rely on indirect estimates of mixture 

toxicity—made from the toxicities of individual 

pesticides that occur—primarily because toxicity data are seldom available for specific unique 

mixtures that occur in the environment. The large 

number of unique mixtures present in streams, 

and to a lesser extent in ground water, make it 

impractical to assess the potential effects of all 

that are encountered (Lydy and others, 2004). 

NAWQA results provide an assessment of the 

unique mixtures that were detected most frequently—such as those summarized in tables 

5–1 and 5–2—and make it possible to prioritize 

specific mixtures for further investigation. In 

developing a strategy for investigation, however, it must be kept in mind that findings about 

mixtures, like those regarding individual pesticides, are limited to the pesticides measured, and 

are influenced by the analytical methods used in 

this study. Thus, NAWQA data yield a minimum 

assessment of the occurrence of mixtures because 

of the limited number of pesticides and degradates that were analyzed.

NAWQA data on mixtures are beginning to 

be used to prioritize toxicological investigations. 

For example, the ATSDR is in the process of 

evaluating the toxicity of the mixture of atrazine, 

deethylatrazine, diazinon, nitrate, and simazine 

(ATSDR, 2004a) because of the frequency of its 

occurrence in public-supply and domestic wells 

that were sampled by NAWQA (Squillace and 

others, 2002). The importance to aquatic life, 

wildlife, and humans of mixtures that occur in 

streams and ground water is difficult to determine, and will require multiple approaches over 

an extended period of time.

Pesticide Degradates

Once released into the environment, pesticides are transformed over time by a variety of 

chemical, photochemical, and biologically mediated reactions into other compounds, which are 

referred to in this report as degradates. With time, 

degradates may become as prevalent as parent 

pesticides—or more so—depending on their rate 

of formation and their relative persistence. For 

example, deethylatrazine, which is a degradate of 

atrazine and other triazine herbicides, was one of 

the few degradates included in routine NAWQA 

analyses, one of the most frequently detected 

pesticide compounds in water (fig. 4–2), and one 

of the most frequent contributors to pesticide 

mixtures (fig. 5–8). In addition, degradates and 

by-products of organochlorine pesticides were 

among the most commonly detected pesticide 

compounds in fish (fig. 4–4). This and other 

evidence from many studies in a wide range of 

settings indicate that a diverse range of pesticide 

compounds routinely occur along with mixtures 

of parent pesticides (Boxall and others, 2004).

Degradates, like their parent compounds, 

have the potential to adversely affect water 

quality, depending on their toxicity. Degradates 

may be either more or less toxic than their parent pesticides, although most have toxicities to 

aquatic life that are similar to, or lower than, 

those of their parent compounds (Sinclair and 

Boxall, 2003; see accompanying sidebar, p. 81). 

For some pesticides that have not been registered 

or reregistered by USEPA during the last several 

years, the toxicities of degradates have not been 

evaluated, but current registration requirements 

include assessment of the toxicities of major 

degradates, as described in the accompanying 

sidebar on USEPA risk assessments (p. 86).

The rates of pesticide transformation and 

degradate formation vary widely among pesticides and under different environmental conditions, as discussed in Chapter 2. Each transformation reaction requires specific physical, 

chemical, and biological conditions. For example, most oxidation reactions require the presence 

of dissolved oxygen, whereas reduction reactions 

require its absence. Photochemical reactions 

require the presence of sunlight that has sufficient energy to break specific chemical bonds. 

Many transformations—such as the conversion 

of atrazine to deethylatrazine, or the formation of 

alachlor ethanesulfonic acid (ESA) from alachlor 

in soil—will not occur without the assistance of 

microbes or other organisms (Barbash, 2004). 

Selected transformations of atrazine are dis80 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Recent advances in analytical methodology and greater access 

to analytical standards have led to the detection of degradates from 

a wide variety of pesticides and other compounds in surface water, 

ground water, precipitation, air, and sediment (Boxall and others, 

2004). Many of these degradates are more persistent in the environment than their parent compounds, and many are more mobile, as 

well.

In most cases, degradates have similar or lower toxicity to aquatic 

life than their parents, but some are more toxic. In a recent review 

of available ecotoxicity data for degradates of pesticides and other 

compounds, Sinclair and Boxall (2003) reported that 41 percent of 

degradates were less toxic than their parent compounds and 39 

percent had a toxicity similar to their parents; however, 20 percent 

were more than 3 times more toxic than their parent compound and 

9 percent were more than 10 times more toxic. In general, the greatest increases in toxicity from parent to degradate were observed for 

parent compounds that had a low toxicity. Similar patterns are apparent for degradates of 8 pesticides frequently detected by NAWQA 

(fig. 5–11), with 28 percent of the degradates being more toxic to fish 

than their parent compounds and 21 percent being more toxic to 

daphnids. 

Because pesticides and their degradates are more commonly 

detected in environmental media as mixtures than as isolated compounds, assessments of their potential biological effects should 

account for the combined effects of multiple substances. As discussed in further detail in Chapter 6 of this report, the influence of a 

given compound A on the toxicity of a second compound B may be 

antagonistic (overall toxicity less than that of A and B combined), 

additive (overall toxicity is roughly equal to that of A and B combined) 

or synergistic (overall toxicity exceeds the added effects of the two 

compounds). The “mixture risk quotient” can be used to assess the 

combined risk of multiple compounds (parent compounds and/or 

degradates) simultaneously, based on the assumption that the toxic 

effects of different compounds are additive. This approach was 

recently applied by Fenner and others (2002) to assess the potential 

toxicities of predicted concentrations of nonylphenol ethoxylate 

(NPEO, a widely used nonionic surfactant) and its degradates to 

aquatic biota in Switzerland. 

Although a variety of methods have been used to estimate mixture 

risk quotients, Fenner and others (2002) computed this parameter by 

summing the ratios of the concentrations predicted in Swiss rivers 

to the no-effect levels (for acute health effects in aquatic organisms) predicted for individual compounds in hypothetical mixtures 

of NPEO and its degradates. Risk quotients for NPEO alone and for 

each of the individual NPEO degradates were all below 1 for water 

and sediment, indicating relatively low risk to aquatic ecosystems 

with respect to acute (but not necessarily chronic) effects. The risk 

quotient calculated for the mixture of NPEO plus all of its transformation products, however, was 2.2, indicating a high risk of acute health 

effects for aquatic organisms if the toxicities are additive. This and 

other studies indicate that, in some instances, degradates from pesticides and other anthropogenic compounds may be of concern in the 

environment. An improved understanding is therefore needed of the 

environmental distributions, patterns of co-occurrence, and toxicities 

of these compounds in the hydrologic system—both in isolation and 

in mixtures.

0.1

0.1 1 10 100 1,000 10,000 100,000 1,000,000

1

10

100

1,000

10,000

100,000

1,000,000

2,4-D

Alachlor

Atrazine

Diuron

EPTC

Chlorpyrifos

Diazinon

Malathion

Daphnids Fish

Herbicides

Insecticides

LC50 or EC50 for degradates, in micrograms per liter

LC50 or EC50 for parent pesticides, in micrograms per liter

degradate is 10 times less toxic

degradate is 10 times more toxic

Comparison of parent and degradate toxicities

Figure 5–11. Comparison of acute toxicity values (LC50 or EC50; see glossary) for parent compounds 

and degradates for 8 pesticides frequently detected by NAWQA in water samples from streams 

shows that 28 percent were more toxic to fish and 21 percent of the degradates were more toxic to 

daphnids than the parent pesticides. Multiples of the same symbol are different degradates of the 

same parent pesticide.

Potential Risks of Pesticide Degradates to Aquatic Life

Alistair B. A. Boxall, University of York, United Kingdom

Chris J. Sinclair, Central Science Laboratory, York, United Kingdom

Dana W. Kolpin, U.S. Geological Survey

Complexities: Seasonality, Mixtures, and Degradates 81

Chapter 5
N

N

N

N N

Cl

H H

N

N

N

N N

Cl

H H

N

N

N

N N

H H

atrazine

deisopropyl atrazine deethylatrazine hydroxyatrazine

H

N

N

N

N N

Cl

H H

H

OH

3CO2 2CO2

H O2

HCl

Dealkylation

(microbes required)

Dealkylation

(microbes required)

Hydrolysis

Atrazine transformation

Figure 5–12. Atrazine transforms to three primary degradates (although there are others as well), one of which—

deethylatrazine—was routinely measured by NAWQA. Two of these transformation reactions require microbes, resulting 

in the formation of deethylatrazine and deisopropyl atrazine. The third is hydrolysis, an abiotic reaction with water that 

produces the degradate hydroxyatrazine.

played in figure 5–12, to show three of the multiple pathways that these reactions may involve, 

and to provide a sense of the complexity involved 

with tracking the formation, transport and fate of 

degradates for all of the pesticides in use. This 

example also illustrates the varying effects of 

different reactions on toxicity, discussed earlier. 

The first two reactions produce degradates with 

mammalian toxicities similar to that of the parent compound (atrazine). By contrast, the third 

reaction generates a compound (hydroxyatrazine) 

with substantially lower mammalian toxicity, 

owing to the removal of chlorine (Jordan and others, 1970; Rodriguez and Harkin, 1997). 

As noted earlier, degradates are often more 

prevalent than their parent pesticides in streams 

and ground water, particularly when conditions favor transformation to degradates that are 

chemically persistent. In parts of some hydrologic systems, the concentrations of degradates 

may exceed those of the parent pesticides 

throughout much of the year. In surface waters, 

degradates often predominate when much of 

the streamflow is either from ground water, or 

from surface runoff occurring long enough after 

pesticide applications for the parent pesticide to 

have substantially transformed. For example, the 

summed concentrations of atrazine, cyanazine, 

acetochlor, alachlor, and metolachlor in the Iowa 

River in the Eastern Iowa Basins changed rapidly 

in response to the timing of their applications, but 

the summed concentrations of their degradates 

were higher and relatively constant throughout 

most of the year (fig. 5–13). Similarly, in the 

Mermentau River in the Acadian–Pontchartrain 

Drainages, the concentration of the insecticide 

fipronil reached its maximum value immediately 

following the spring application season, and then 

declined, to be exceeded by concentrations of 

fipronil degradates from June to February. This 

cycle repeated itself with the springtime applications the following year (fig. 5–14). Information 

on the concentrations and fluxes of degradates—

especially in relation to those of their parent 

82 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 5

1996 1997 1998

January

March

May

July

September

November

January

March

May

July

September

November

January

March

May

July

September

November

100

10

1

0.1

100,000

10,000

1,000

100

10

Concentration,

in micrograms per liter

Streamflow, 

in cubic feet per second

5_13

Streamflow

Total

degradates

Herbicides and their degradates in the Iowa River, Iowa

Total

parent

herbicides

Figure 5–13. The summed concentrations of the parent herbicides atrazine, 

cyanazine, acetochlor, alachlor, and metolachlor in the Iowa River (Eastern Iowa 

Basins) rose and fell in response to spring applications, whereas the summed 

concentrations of their degradates remained relatively steady and at higher levels 

throughout most of the year (Schnoebelen and others, 2003).

2000 2001

Fipronil

Desulfinylfipronil

Fipronil sulfide

Fipronil sulfone

0.01

0.001

0.1

1

10

Concentration,

in micrograms per liter

March

April

May

June

July

August

January

September

October

NovemberDecember

April

March

February

5_14

Fipronil and its degradates in the Mermentau River, Louisiana

Figure 5–14. Concentrations of the insecticide fipronil and its degradates in the Mermentau 

River (Acadian–Pontchartrain Drainages) peaked in March or April, following applications. 

Although fipronil dominated the total concentration of fipronil compounds during the high-use 

period, concentrations of its degradates were greater during the rest of the year (Demcheck and 

others, 2004).

Complexities: Seasonality, Mixtures, and Degradates 83
Deethylatrazine-to-Atrazine Ratios

As pesticides are transported through 

the hydrologic system, transformations occur 

continuously and at various rates, resulting in 

changes in the proportional relations between 

parent pesticides and their degradates with time 

and space. In general, the extent of transformation increases with increasing residence time in 

the hydrologic system. As a result, degradate-toparent compound concentration ratios—such as 

the deethylatrazine-to-atrazine ratio (Adams and 

Thurman, 1991)—have been used as indicators 

of residence time in the environment. Because 

the transformation of atrazine to deethylatrazine 

requires microbial assistance (fig. 5–12)—and 

microbial populations are generally much higher 

in the soil than at greater depths beneath the 

land surface or in surface waters—the deethylatrazine-to-atrazine ratio provides an indication 

of the amount of time that atrazine has been in 

contact with soil. In streams, the deethylatrazineto-atrazine ratio increased with the time elapsed 

between atrazine applications and sampling—

from the lowest values during atrazine applications in the spring, to higher values in autumn, 

and to the highest values in winter, just before 

applications (fig. 5–16). Figure 5–17 shows how 

the ratio changed over time during the year in 

the White River. Because the analytical recovery 

(the proportion of the actual total concentration 

that could be measured) for deethylatrazine was 

lower than for atrazine, the ratios reported here 

underestimate the true value, but the focus of 

this analysis is on the relative magnitudes of the 

ratios among different media, settings, and times 

of year, rather than their absolute values.

Deethylatrazine-to-atrazine ratios were 

generally higher in ground water than in streams 

throughout the year, reflecting the longer periods 

of time spent in contact with soil for the atrazine 

detected in the ground-water system, relative to 

streams (fig. 5–16). The proportions of deethylatrazine in water collected from major aquifers—which generally represents ground water 

that is deeper and older than water collected from 

shallow wells—were typically higher than those 

measured in the shallow ground water sampled 

within agricultural areas. 

Figure 5–15. In ground water of the Upper Illinois 

River Basin, degradates (light bars) were generally 

detected more frequently than parent compounds 

(dark bars) for the acetanilide herbicides (acetochlor, alachlor, and metolachlor), but at comparable 

frequencies for atrazine (Groschen and others, 

2004). 

Frequency of detection, as percentage of

wells with 0.05 micrograms per liter or greater

Acetochlor not detected 

Metolachlor

Alachlor

Atrazine

Acetochlor ESA

Acetochlor OA

Metolachlor OA

Metolachlor ESA

Alachlor ESA

Alachlor OA

Deethyldeisopropyl atrazine

Deisopropyl atrazine

Hydroxyatrazine

Deethylhydroxyatrazine

Deisopropylhydroxyatrazine

0 10 20 30 40

5_15

Herbicide degradates in shallow

ground water in agricultural areas

of the Upper Illinois River Basin

pesticides—contributes to our understanding of 

the environmental fate of pesticides as they move 

and transform within the hydrologic system. 

 In ground water, degradates were often 

detected more frequently, or at higher concentrations, than their parent pesticides. For example, 

in ground water of the Upper Illinois River Basin, 

degradates of acetochlor, alachlor, metolachlor, 

and atrazine accounted for substantially more 

detections than the parent compounds (fig. 5–15). 

Two of the principal factors likely to be responsible for this general observation are that (1) 

ground water recharges through soil where 

microbial populations—and thus transformation 

rates—are relatively high, and (2) residence times 

in ground water are usually long prior to sample 

collection, allowing more time for transformations to occur than is usually the case for surface 

waters. 

84 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 5

Figure 5–16.  The ratios of deethylatrazine to

atrazine concentrations—which increase with the

transformation of atrazine to deethylatrazine—

increased with increasing residence time in the

hydrologic system. (Ratios were computed from

μg/L concentrations for all NAWQA stream and

ground-water samples in which both compounds

were detected.)

Figure 5–17.  The ratios of deethylatrazine to atrazine

concentrations in the White River (White River Basin)

were lowest in the spring, following widespread atrazine

applications. The ratio generally increased through the summer

and winter as atrazine transformed to deethylatrazine and

ground water made up an increasing proportion of streamflow.

(Ratios were computed from μg/L concentrations; modified

from Carter and others, 1995.)

100

50

75

25

0

0 1 2 3 4

Frequency that ratio

was equaled or exceeded,

as a percentage of sites or studies

Deethylatrazine/atrazine ratio

Preapplication (Dec.–March)

Application (April–July)

Post application (Aug.–Nov.)

Agricultural land use

Major aquifers

Streams

Ground Water

5_16

National overview of deethylatrazine/atrazine 

ratio in streams and ground water May June July August October November April May

June

July

August

September

December

January

February

March

1991 1992

0.2

0.4

0.6

0.8

0.0 0

5,000

10,000

20,000

30,000

40,000

15,000

25,000

35,000

45,000

Deethylatrazine/atrazine ratio

Streamflow, in cubic feet per second 

Streamflow

Deethylatrazine/atrazine ratio

Deethylatrazine/atrazine ratio in the

White River, Indiana

Importance of Pesticide Degradates

NAWQA results are consistent with findings 

from other studies that found that pesticide degradates occur frequently in streams and ground 

water (Battaglin and others, 2001; Scribner and 

others, 2003; Kolpin and others, 1998; Kolpin 

and others, 2004). Assessment of the occurrence, 

distribution, and toxicities of pesticide degradates 

in the hydrologic system is important because 

of the potential effects of these compounds on 

human health and the environment (Sinclair and 

Boxall, 2003), as well as their value for understanding the ultimate fate of pesticides in the 

hydrologic system (Barbash, 2004). Pesticide 

degradates should continue to be considered 

and accounted for in assessments of pesticide 

exposure and in evaluating the potential effects 

of pesticides. Improved assessment of pesticide 

degradates will require expanded coverage of 

degradates in water-quality monitoring, continued research on pesticide transformations and 

transport, and continued attention to these compounds in toxicity studies, including as components of pesticide mixtures (see Chapter 6).

White River, Indiana.

Complexities: Seasonality, Mixtures, and Degradates 85
Federal Insecticide, Fungicide, and Rodenticide 

Act (FIFRA)

When USEPA characterizes the risks of a pesticide to humans and 

the environment to meet the requirements of FIFRA, the agency evaluates both the parent pesticide and its degradates. Before a pesticide 

is registered, USEPA reviews and evaluates available studies on the 

pesticide’s properties and effects, including its degradates (USEPA, 

2005a). The types of data needed vary depending on how and where 

the pesticide is used (USEPA, 2004a). 

Available studies may provide information on the degradation 

rates, mobility parameters, and toxicity for potentially important degradates. If adequate data are not available for specific degradates, 

USEPA’s risk assessments for drinking-water exposure under FIFRA 

assume that degradates are highly persistent and mobile. If different 

toxicities are expected between degradates and the parent pesticide, concentrations are estimated separately. If the same toxicity 

is expected, concentrations of the parent and degradates are added 

together. These risk estimates are often conservative (protective) 

because it is frequently assumed that the degradate and parent have 

equal toxicities and that they are mobile and persistent. If data are 

available for a specific degradate, however, then those data are used. 

In some cases, degradates are known to be more toxic than the parent compound. In situations where degradates occur in substantial 

amounts or are of toxicological concern, risk assessments include a 

quantitative or qualitative analysis of the potential implications of an 

organism’s exposure to these degradates, in addition to the parent 

pesticide. 

USEPA environmental fate scientists work with human health and 

ecotoxicology scientists to identify the degradates of toxicological 

concern. The formation of degradates is monitored as part of fate 

studies required for pesticide registration. Methods are used that 

have detection limits sufficiently low to allow for detailed tracking 

of the production of degradates. Degradates formed at greater than 

10 percent of radioactively labeled parent pesticide are considered 

major degradates and must be identified (USEPA, 1982). The 10-percent criterion is a general guideline, such that degradates approaching concentrations of 10 percent of the parent are usually identified as 

well. In addition, degradates of known toxicological or ecotoxicological concern must be quantified and identified even when present at 

lower levels. 

When environmental monitoring data are available for pesticide 

degradates, the data are characterized and summarized in USEPA’s 

assessments for FIFRA. In evaluating monitoring data, scientists 

evaluate the analytical methods used, pesticide-use information, and 

the design of the monitoring studies. Monitoring data on the occurrence of degradates are included in the FIFRA risk assessment, but 

the assessment of risk also depends on a variety of additional factors, 

including the mode of toxicity of the degradate—information that is 

needed to determine if concentrations of the parent pesticide and 

degradates can reasonably be aggregated to assess risk.

Clean Water Act (CWA)

Ambient water-quality criteria, developed by USEPA under section 304(a) of the CWA, focus on individual chemicals. If a degradate 

is toxicologically important, a separate criterion may be developed 

for the degradate. Human health ambient water-quality criteria exist 

for DDT, DDE, and DDD; endrin and endrin aldehyde; heptachlor and 

heptachlor epoxide; and endosulfan and endosulfan sulfate. Wholeeffluent toxicity tests, described in Chapter 6 in relation to assessing 

potential effects of pesticide mixtures on aquatic life, also provide an 

approach for assessing degradate toxicity. 

Safe Drinking Water Act (SDWA)

Maximum Contaminant Levels (MCLs) are legally enforceable 

drinking-water standards developed by USEPA under the SDWA. 

Although drinking-water standards have typically been developed 

only for the pesticide parent, the SDWA does not preclude USEPA 

from developing standards for pesticide degradates. Several unregulated pesticide degradates are listed on USEPA’s drinking-water 

Contaminant Candidate List (CCL) and its Unregulated Contaminant 

Monitoring Regulation (UCMR) (USEPA, 2005b,c). Once sufficient 

information and data are available on health risks, occurrence, analytical methods, and treatment technologies, USEPA will determine 

whether any of the listed pesticide degradates are candidates for 

future drinking-water standards. 

USEPA also develops drinking-water Health Advisories for chemical substances, including some pesticides and pesticide degradates. 

Health Advisories, which are not legally enforceable, provide technical guidance for Federal, State, Tribal and local officials in the event 

of an emergency spill or contamination situation. USEPA periodically 

updates Health Advisories when new information becomes available 

(USEPA, 2005d).

86 Pesticides in the Nation’s Streams and Ground Water, 1992–2001

Pesticide Degradates in USEPA Risk Assessments and Regulations
Potential for Effects on Human Health, Aquatic Life, and Wildlife 87

Chapter 6

The occurrence of pesticides in streams and ground water 

raises the question—Do pesticides occur at concentrations 

that may affect human health or stream ecosystems? Comparisons of concentrations measured by NAWQA to water-quality 

benchmarks provide a screening-level assessment of the potential for adverse effects. Concentrations of pesticides detected 

in streams and wells were usually lower than human-health 

benchmarks, indicating that the potential for effects on drinking-water sources probably is limited to a small proportion of 

source waters. More than half of the wells sampled, but none 

of the stream sites that were sampled, are current sources of 

drinking water. Concentrations in streams more frequently 

exceeded benchmarks for aquatic life and fish-eating wildlife.

More than half of all agricultural streams sampled and more 

than three-quarters of all urban streams had concentrations of 

pesticides in water that exceeded one or more benchmarks for 

aquatic life. In addition, organochlorine pesticides—most uses 

of which were discontinued 15–30 years ago—still exceeded 

benchmarks for aquatic life and fish-eating wildlife in bedsediment or fish-tissue samples from many streams.

This chapter examines the 

potential for pesticides to have 

adverse effects on human 

health, aquatic life, and fisheating wildlife. The potential 

is assessed by comparing 

measured concentrations with 

water-quality benchmarks. 

The screening-level 

assessment provides 

indications of the distribution 

of potential effects and 

the pesticides that may 

cause them, which can be 

used to prioritize further 

investigations.

6

Potential for Effects on Human Health, 

Aquatic Life, and Wildlife
Screening-Level Assessment of 

Potential Effects 

The potential for pesticide concentrations 

measured by NAWQA to adversely affect human 

health, aquatic life, or fish-eating wildlife was 

evaluated by screening-level assessments similar 

in concept to USEPA screening-level assessments 

(USEPA, 2004d). The NAWQA screening-level 

assessments compare site-specific estimates of 

pesticide exposure (concentration statistics or 

concentrations determined from measurements of 

pesticides in various media at NAWQA sampling 

sites) with water-quality benchmarks derived 

from standards and guidelines established by 

USEPA, toxicity values from USEPA pesticide 

risk assessments, and selected guidelines from 

other sources. The characteristics and limitations 

of screening-level assessments are summarized 

in the accompanying sidebar on page 89. The 

USEPA standards, guidelines, and toxicity values 

were developed by USEPA as part of the Federal 

process for assessing and regulating pesticides, as 

summarized in the sidebar on page 90. 

NAWQA studies were not designed to 

evaluate specific effects of pesticides on humans, 

aquatic life, or fish-eating wildlife. The screening-level assessment is not a substitute for either 

risk assessments, which include many more 

factors (such as additional avenues of exposure), 

or site-specific studies of effects. Rather, comparisons of measured concentrations with waterquality benchmarks provide a perspective on the 

potential for adverse effects, as well as a framework for prioritizing additional investigations 

that may be warranted. Measured concentrations 

that exceed a benchmark do not necessarily 

indicate that adverse effects are occurring—they 

indicate that adverse effects may occur and that 

sites where benchmarks are exceeded may merit 

further investigation. 

Screening-level assessments should be 

considered as a first step toward addressing the 

question of whether or not pesticides are present 

at concentrations that may affect human health, 

aquatic life, or wildlife. They provide a perspective on where effects are most likely to occur and 

what pesticides or degradates may be responsible. As improved data on toxicity and environmental concentrations are developed, benchmarks 

and exposure estimates can be updated, and the 

assessments can be improved and expanded. 

USGS works closely with USEPA to assist them 

with incorporating NAWQA findings into their 

risk assessments. 

NAWQA screening-level assessments for 

pesticides are presented in this chapter for human 

health (concentrations in water), aquatic life 

(concentrations in water and bed sediment), and 

fish-eating wildlife (concentrations in whole 

fish). The selection of benchmarks for each of 

these assessments is described below along with 

results and the specific values and sources for 

benchmarks used are provided in Appendix 3. 

Each type of benchmark selected for use in the 

screening-level assessment applies to a specific 

sampling medium (such as water or bed sediment) and to a specific use of the water resource 

(such as for drinking water or to support aquatic 

life). Priority was given to (1) benchmarks based 

on USEPA standards, guidelines, or toxicity values; (2) benchmarks that are nationally relevant 

because of the nature or breadth of toxicity data 

on which they are based; and (3) systematically 

derived suites of benchmarks that share a common methodology and are available for a large 

number of NAWQA analytes.

88 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 6

Characteristics and Limitations of the Screening-Level Assessment 

of Potential Effects 

The NAWQA screening-level assessment provides an initial perspective on the potential importance of 

pesticides to water quality in a national context by comparing measured concentrations with water-quality 

benchmarks. The screening-level assessment is not a substitute for risk assessment, which includes many 

more factors, such as additional avenues of exposure. The screening-level results are primarily intended to 

identify and prioritize needs for further investigation and have the following characteristics and limitations. 

Most benchmarks used in this report are estimates of no-effect levels, such that concentrations below 

the benchmarks are expected to have a low likelihood of adverse effects and concentrations above a 

benchmark have a greater likelihood of adverse effects, which generally increases with concentration. 

The presence of pesticides in streams or ground water at concentrations that exceed benchmarks does 

not indicate that adverse effects are certain to occur. Conversely, concentrations that are below benchmarks do not guarantee that adverse effects will not occur, but indicate that they are expected to be 

negligible (subject to limitations of measurements and benchmarks described below).

The potential for adverse effects of pesticides on humans, aquatic life, and fish-eating wildlife can only 

be partially addressed by NAWQA studies because chemical analyses did not include all pesticides and 

degradates. In addition, some compounds analyzed by NAWQA do not have benchmarks.

Most benchmarks used in this report are based on toxicity tests of individual chemicals, whereas 

NAWQA results indicate that pesticides usually occur as mixtures. Comparisons to single-compound 

benchmarks may tend to underestimate the potential for adverse effects. 

Water-quality benchmarks for different pesticides and media are not always comparable because they 

have been derived by a number of different approaches, using a variety of types of toxicity values and 

test species.

For some benchmarks, there is substantial uncertainty in underlying estimates of no-effect levels, 

depending on the methods used to derive them and the quantity and types of toxicity information on 

which they are based. This is especially true of fish-tissue benchmarks for the protection of fish-eating 

wildlife, for which there is no consensus on national-scale benchmarks or toxicity values.

Estimates of pesticide exposure derived from NAWQA concentration measurements are also uncertain—particularly estimates of short-term exposure of aquatic organisms to pesticides in stream water. 

Generally, short-term average concentrations in stream water, such as 4-day values, are underestimated from NAWQA data. 

•

•

•

•

•

•

•

Screening-Level Assessment for 

Human Health 

NAWQA studies, as emphasized in 

Chapter 3, characterized the quality of untreated 

water from streams and ground water, whether 

or not that water was used as a source of drinking water during the study period. More than half 

of the wells sampled for ground-water studies, 

but none of the stream sites that were sampled, 

were sources of domestic or public water supplies. In this report, measured concentrations 

of pesticides in all wells and streams sampled, 

whether or not they were sources of drinking 

water during the study period, are compared with 

human-health benchmarks derived from available 

USEPA drinking-water standards and guidelines 

as a starting point for understanding the potential importance of pesticides in a human-health 

context. The benchmarks are described in the 

accompanying sidebar on page 91 and values are 

listed in Appendix 3A. 

Comparisons of human-health benchmarks 

to the concentrations observed in NAWQA 

studies provide a perspective on the potential 

importance to human health as use of water 

resources expands, but they are not appropriate 

for assessing current compliance with drinking-water regulations. A measured concentration 

or computed annual mean that is greater than a 

benchmark indicates the potential need for further investigation if such water either is presently 

used as a drinking-water source, or may be used 

as a source in the future. A concentration greater 

than a Maximum Contaminant Level (MCL), 

even in water that is now a source of drinking 

water, does not indicate violation of a standard. 

For water currently used as a drinking-water 

source, pesticide concentrations in finished water 

may be lower than those measured in untreated 

Potential for Effects on Human Health, Aquatic Life, and Wildlife 89
The potential effects of pesticides on humans and the environment 

are managed under several Federal Acts and regulated through a 

combination of Federal, State, and Tribal responsibilities. The Federal 

Insecticide, Fungicide, and Rodenticide Act (FIFRA), the Federal Food, 

Drug and Cosmetic Act (FFDCA), the Food Quality Protection Act 

(FQPA), the Safe Drinking Water Act (SDWA), and the Clean Water Act 

(CWA)—all of which are administered by USEPA and partner agencies—provide the regulatory framework that affects the assessment 

and control of pesticides and their degradates in water resources.

The FIFRA, first enacted in 1947 and amended most recently by 

the FQPA in 1996, provides the original framework for the Federal 

pesticide licensing program administered by USEPA and covers the 

evaluation and registration of pesticides for specific uses. Before a 

pesticide may be registered it must be shown, among other things, 

that it will not generally cause “unreasonable adverse effects” 

to water, air, land, plants, and man and other animals. The FFDCA 

authorizes USEPA to set maximum residue levels, or tolerances, for 

pesticides used in or on foods or animal feed, and mandates strong 

provisions to protect infants and children. Before a pesticide registration may be granted for use on a food commodity, a tolerance must be 

set or an exemption from a tolerance granted. The FQPA amended the 

FIFRA and FFDCA to set more stringent safety standards for new and 

old pesticides. Among its provisions are that (1) human-health assessments must consider aggregate exposures, including all dietary, 

drinking-water, and nonoccupational exposure; and (2) assessments 

leading to tolerance decisions must consider, relative to human 

health, the cumulative effects and common mode of toxicity among 

related pesticides, the potential for endocrine-disrupting effects, and 

appropriate safety factors to further protect infants and children.

Through application of the FIFRA, FFDCA, and FQPA, USEPA determines the specific conditions under which a pesticide can be legally 

sold, distributed, and used in the United States, including where, how, 

and at what application rates pesticides may be used.

The SDWA was originally passed by Congress in 1974 to protect 

public health by regulating the Nation’s public drinking-water supply. The law was amended in 1986 and 1996 and requires protection 

of drinking water and its sources, including rivers, lakes, reservoirs, 

springs, and wells (the SDWA does not regulate private wells that 

serve fewer than 25 individuals). The SDWA authorizes the USEPA 

to set national health-based standards for drinking water to protect 

against both naturally occurring and manmade contaminants that may 

be found in drinking water. The USEPA works with States and water 

utilities to make sure that these standards are met.

 The CWA (originally the Federal Water Pollution Control Act 

Amendments of 1972, and subsequently amended several times) 

provides for protection against releases of toxic chemicals. Section 101(a)(3) of the CWA states that “it is national policy that the 

discharge of toxic pollutants in toxic amounts be prohibited.” Section 

303(c) of the CWA requires States to develop water-quality standards 

to protect the public health or welfare, enhance the quality of water, 

and serve the purposes of the CWA. The control of the discharge 

of toxic substances is a key objective of the National Pollutant Discharge Elimination System (NPDES) and water-quality standards 

programs. Section 304(a) of the CWA requires USEPA to develop 

and publish and, from time to time, revise ambient water-quality criteria for the protection of both human health and aquatic life. When 

final, these criteria provide USEPA’s recommendations to States and 

authorized Tribes as they develop their own water-quality standards. 

USEPA-recommended criteria are not regulations, and they do not 

impose legally binding requirements on USEPA, States, authorized 

Tribes, or the regulated community. However, USEPA-recommended 

criteria may form the basis for State or Tribal water-quality standards 

and become enforceable through NPDES permits or other environmental programs. USEPA’s role in this process, in addition to providing 

criteria recommendations, is to review and approve the water-quality 

standards developed by States and Tribes. 

90 Pesticides in the Nation’s Streams and Ground Water, 1992–2001

Federal Regulation of Pesticides in Water 
Chapter 6

of the year and concentrations would thus be 

compared with the 37.5 μg/L benchmark. Of the 

3 sites with at least weekly sampling, 1 exceeded 

the 37.5 μg/L benchmark. Use of benchmarks 

from the new risk analysis would, therefore, 

result in screening-level exceedances for 2 sites 

instead of 5 sites, although inclusion of the other 

chlorinated degradates could increase the number 

of sites with exceedances. Further analysis of the 

distribution of atrazine concentrations in streams 

nationwide is presented in Chapter 7.

Human-Health Benchmarks for 

Pesticides in Water

Benchmarks for assessing the potential for pesticides in water to affect 

human health were derived from three types of USEPA drinking-water 

standards and guidelines developed by USEPA’s Office of Water (USEPA, 

2004c, 2005e). One or more drinking-water standards or guidelines are 

available for 47 of the 83 pesticides and degradates analyzed by NAWQA 

(Appendix 3A). 

Maximum Contaminant Level (MCL)—The maximum permissible concentration of a contaminant in water that is delivered to any user of a public water 

system. This is an enforceable standard issued by USEPA under the Safe 

Drinking Water Act and established on the basis of health effects and other 

factors (analytical and treatment technologies, and cost).

Lifetime Health Advisory (HA-L)—The concentration of a chemical in drinking water that is not expected to cause any adverse, noncarcinogenic effects 

over a lifetime exposure. A health advisory is not a legally enforceable Federal 

standard, but serves as technical guidance to assist Federal, State, Tribal 

and local officials. The HA-L is based on toxicity (dose-response) information 

for the chemical. It assumes lifetime consumption of 2 liters (L) of water per 

day by a 70-kilogram (kg) adult, and that 20 percent of total exposure to the 

contaminant comes from drinking water (80 percent is assumed to come from 

other sources). 

10-6 Cancer Risk Concentration—The concentration of a chemical in drinking water corresponding to an excess estimated lifetime cancer risk of 1 in 1 

million (10-6). These values are calculated from the estimated cancer potency, 

which is derived using a conservative (protective) model of carcinogenesis, 

so that the cancer risk is an upper-limit estimate. The definition of “acceptable” level of cancer risk is a policy issue, not a scientific one. USEPA reviews 

individual State and Tribal policies on cancer risk levels as part of its oversight of water-quality standards under the Clean Water Act. USEPA’s policy is 

to accept measures adopted by States to limit cancer risk to the range of 10-6

to 10-4 (USEPA, 1992a). The concentration corresponding to a cancer risk of 

10-6 was used as the benchmark for the NAWQA screening-level assessment, 

consistent with the conservative (protective) nature of such assessments.

Application of Human-Health Benchmarks for Water

If available, the MCL was used as the human-health benchmark for a given 

pesticide. For pesticides with no MCL, the lower of the HA-L and the 10-6 cancer risk concentration was used. Human-health benchmarks were compared 

with time-weighted annual mean concentrations of pesticides in streams, as 

well as with concentrations measured in individual wells for ground water. 

water (depending on whether and how the water 

is treated), because some drinking-water treatment processes reduce pesticide concentrations. 

In addition, NAWQA sampling methods were 

not designed to meet the specific sampling and 

analytical requirements for determining compliance with an MCL. 

Streams

Annual mean concentrations of pesticides 

in the 186 streams sampled by NAWQA were 

seldom greater than human-health benchmarks 

during 1992–2001, and most exceedances were 

in streams draining agricultural and urban watersheds (fig. 6–1). Specifically, pesticide concentrations exceeded one or more human-health 

benchmarks in about 10 percent of agricultural 

streams, 7 percent of urban streams, and in 1 

of the 65 mixed-land-use streams sampled by 

NAWQA. No benchmarks were exceeded in the 

eight undeveloped streams that were sampled. 

The streams sampled by NAWQA that 

had concentrations of a pesticide greater than 

a human-health benchmark were clustered in a 

few regions. Specifically, 6 agricultural streams 

and 1 mixed-land-use stream with concentrations greater than one or more benchmarks (5 of 

7 streams for atrazine and 4 of 7 for cyanazine) 

were in the Corn Belt or southern Mississippi 

River Basin, where atrazine and cyanazine use 

was high during the study period (fig. 6–2). Two 

agricultural streams, 1 in California and 1 in 

Washington, had concentrations of dieldrin that 

were greater than its benchmark. The 2 urban 

streams in which benchmarks were exceeded are 

in Texas (diazinon) and Hawaii (dieldrin).

A new analysis of atrazine’s potential 

risks has been developed by USEPA as part 

of the reregistration process (USEPA, 2003a). 

The analysis is based on the concentrations of 

atrazine and three of its chlorinated degradates, 

referred to, collectively, as “total atrazine.” The 

human-health benchmarks from this new analysis 

are 37.5 μg/L for the 90-day moving average if 

the monitoring frequency is at least weekly and 

12.5 μg/L if monitoring is less frequent. Comparison of these benchmark values to 90-day moving 

averages determined from NAWQA data for the 

sum of atrazine and deethylatrazine (NAWQA 

did not measure the 2 other chlorinated degradates) indicates that 4 of the 5 sites that exceeded 

the MCL-based benchmark also had 90-day averages that exceeded the 12.5 μg/L level. Of these 

4 sites, however, 3 had at least weekly sampling 

frequencies during the high-concentration period 

Potential for Effects on Human Health, Aquatic Life, and Wildlife 91
Figure 6–1. Annual mean concentrations of one 

or more pesticides were greater than a humanhealth benchmark in about 10 percent of agricultural 

streams and about 7 percent of urban streams that 

were sampled. No streams draining undeveloped 

watersheds, and only 1 stream with mixed land uses 

in its watershed, had concentrations greater than 

a benchmark. About 1 percent of all domestic and 

public-supply wells sampled had concentrations 

greater than a benchmark. More than half of the 

wells sampled for ground-water studies, but none of 

the stream sites sampled, were sources of domestic 

or public water supplies during the study period. Major aquifers

Stream water

Stream water

Shallow ground water

Shallow ground water

Stream water

Shallow ground water

Stream water

0 25 50 75 100

Percentage of stream sites or wells

with one or more pesticides

exceeding a human-health benchmark

Concentrations greater than 

human-health benchmarks

6_01freqOV

Agricultural areas

Urban areas

Undeveloped areas

Mixed land uses

Figure 6–2. Most streams sampled by NAWQA that had concentrations of a pesticide greater 

than a human-health benchmark were agricultural streams in the Corn Belt and lower Mississippi 

River Basin, where atrazine and cyanazine accounted for the exceedances. Two urban streams 

in Texas and Hawaii had concentrations greater than benchmarks for diazinon and dieldrin, 

respectively. Wells with concentrations greater than benchmarks were widely scattered among 

36 of the 187 ground-water study areas, with the highest proportion in urban areas. Dieldrin 

accounted for most benchmark exceedances in ground water. Streams are indicated if the 

annual mean concentration of one or more pesticides was greater than a benchmark, and 

ground-water studies are indicated if one or more wells had a concentration greater than a 

benchmark. 

Stream sites

Sites or studies with concentrations greater than a human-health benchmark

Ground-water studies

#

#

#

### # # # # # ## # # # # # # # ## ### # # # # # # # # #

# # # ## # ## ## # # ## # # ## # # # ## # # # # ### # # # # # #

# # # # ## ## #

#

#

# #

# # # #

###

# #

#

#

"

#

## #

# ##

# # ## ##

# ## ### ## # #

#

# # # # ### ## # # ## ##

## ###### # # # # # # ### # # # # #

#

#

#

#

Land use

Agricultural

Urban

Undeveloped

Mixed land use or major aquifers

One or more

benchmarks exceeded

No benchmark

exceeded

92 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 6

Although NAWQA findings for streams 

are not directly applicable to drinking-water 

supplies because no NAWQA stream sites were 

located at water-supply intakes, a perspective on potential significance to human health 

is provided by comparing streams sampled by 

NAWQA with streams that serve as sources of 

drinking water and that have similar land uses 

in their watersheds. The Nation’s 1,679 public 

water-supply intakes on streams were classified 

using NAWQA’s land-use definitions (table 3–1 

and fig. 3–1). The stream sites where intakes are 

located are composed of 12 percent agricultural 

streams (194 intakes); 1 percent urban streams 

(22 intakes); 55 percent undeveloped streams 

(926 intakes); and 32 percent streams that drain 

watersheds with mixed land use (537 intakes). 

As a group, however, agricultural streams with 

drinking-water intakes have proportionally less 

agricultural land in their watersheds than do the 

agricultural streams sampled by NAWQA (see 

Chapter 3). Thus, the finding that 10 percent of 

agricultural streams sampled by NAWQA had 

concentrations of pesticides greater than one or 

more benchmarks indicates that probably fewer 

than 10 percent of the 194 drinking-water intakes 

on agricultural streams used source waters 

with concentrations greater than human-health 

benchmarks during the study period. In addition, 

source water may be treated or mixed with other 

water sources to reduce pesticide concentrations 

prior to consumption. 

Overall, the human-health screening-level 

assessment for streams sampled by NAWQA 

during the study period indicates that few of the 

drinking-water intakes that currently withdraw 

water from streams are likely to be located on 

streams with pesticide concentrations greater 

than a benchmark. This broad finding is derived 

from combined data from multiple sites sampled 

in different sampling periods from 1993 to 2000.

In addition, there are sufficient NAWQA stream 

sites with primary sampling years distributed 

throughout the study period to assess changes 

over time in benchmark exceedances for agricultural streams in the corn-and-soybeans crop 

setting (fig. 4–6) and for urban streams. Although 

there were too few exceedances of humanhealth benchmarks at urban sites for meaningful 

assessment of trends, agricultural streams in the 

corn-and-soybeans crop setting had the highest 

frequencies of benchmark exceedances by atrazine and cyanazine. In this agricultural setting, 

the changes in percentages of stream sites that 

had concentrations that exceeded a benchmark 

were different for the two herbicides (fig. 6–3). 

Observations about changes shown in figure 6–3, 

however, are preliminary because they are based 

on different groups of sites for each sampling 

period and site-to-site variability in conditions 

may distort actual trends. There was no clear 

pattern of change through the study period for 

atrazine, but the highest proportion of sites with 

exceedances by atrazine occurred near the end of 

the study period, during 1998–2000. In contrast, 

there was a consistent decrease in exceedances 

for cyanazine during the study period, with none 

during 1998–2000. Data on the agricultural use 

of these two pesticides in the Corn Belt show 

that these changes in frequencies of benchmark exceedances are consistent with their use 

(fig. 6–3). 

Figure 6–3. Changes over time in the percentage of agricultural stream 

sites in the corn-and-soybeans crop setting that had exceedances of 

human-health benchmarks for atrazine and cyanazine generally followed 

trends in use. Sites were grouped according to the year of sampling. 

The 1993–1994 sampling period included 10 sites, the 1995–1997 period 

included 19 sites, and the 1998–2000 period included 6 sites. 1993-1994 1995-1997 1998-2000 1993-1994 1995-1997

1998-2000

Sampling period

0.1

0

0.2

0.3

0.4

1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

25

0

50

75

100

Stream sites with exceedances 

of a benchmark, as a percentage of sites 

sampled during each sampling period

Use intensity, in pounds per acre 

of total agricultural land

Atrazine Cyanazine

Atrazine

Cyanazine

Trends in use and benchmark exceedances for agricultural

streams in the corn-and-soybeans crop setting

Agricultural use of atrazine and cyanazine (IA, IL, IN, MN, NE, OH)

Benchmark exceedances for agricultural streams in the cornand-soybeans crop setting no exceedances

Potential for Effects on Human Health, Aquatic Life, and Wildlife 93
In addition to drinking water, humans also can be exposed to pesticides through consumption of contaminated fish. When persistent, 

hydrophobic compounds, such as organochlorine pesticides, enter a 

stream, they tend to bioaccumulate in fish and other aquatic organisms. Because USEPA sets tolerances only for currently registered 

pesticides, there are no tolerances for the cancelled organochlorine 

pesticides in fish. However, 48 States, the District of Columbia, American Samoa, and three Tribes have issued active fish-consumption 

advisories and safe-eating guidelines to inform people about the 

recommended level of consumption for fish caught in local waters. 

Fish advisories are advice to limit or avoid eating certain fish. USEPA 

has published guidance to States, Territories, Tribes, and local governments to use in establishing fish-consumption advisories (USEPA, 

2000a). As of 2004, there were a total of 79 active fish-consumption 

advisories for chlordane, 67 advisories for DDT and its degradation 

products DDE and DDD, and 22 for dieldrin (USEPA, 2005f,g). Although 

some advisories for organochlorine pesticides have been rescinded 

in recent years, as residues of these pesticides continue to degrade 

slowly in the environment (see chapter 8), new advisories were 

issued in 2004 for DDT, toxaphene, mirex, and chlorinated pesticides 

(USEPA, 2005f). 

USEPA guidelines include recommended screening values, which 

are “concen­trations of target analytes in fish or shellfish tissue that 

are of potential public health concern and that are used as threshold 

values against which levels of contamination in similar tissue collected from the ambient environment can be compared” (USEPA, 

2000a). Screening values were derived separately for carcinogenic 

and noncarcinogenic effects, and USEPA recommends that the lower 

of the two screening values be used for pesticides that have both 

types of effects. USEPA screening values are intended to protect 

the majority of the United States population and are based on average fish and shellfish consumption rates by recreational fishers. For 

potential carcinogens, the recommended screening value is based 

on a maximum acceptable cancer risk of 10-5 (1 in 100,000). USEPA 

screening values are available for 9 of the 12 organochlorine pesticides and pesticide groups (such as total chlordane) measured by 

NAWQA in whole fish.

Comparisons of concentrations of organochlorine pesticide compounds measured in NAWQA fish samples with USEPA screening values are limited in two ways. First, NAWQA analyzes contaminants in 

whole fish, whereas USEPA screening values apply to edible fish tissue. Organochlorine compounds have high affinities for the lipid (fat) 

in fish and other biota. Whole fish generally have higher lipid content 

and, therefore, may have higher organochlorine concentrations than 

the part of the fish that is consumed (fillets). Thus, comparisons of 

NAWQA measurements with USEPA screening values are probably, 

in this sense, worst-case assessments. Second, most fish sampled by 

NAWQA are bottom-feeding species, such as carp and white sucker, 

which are not consumed as frequently as game fish. Depending on 

the compound, however, the difference between game-fish fillets 

and the whole bodies of bottom-feeders may not be significant. For 

example, in a national study of bioaccumulative chemicals in fish, the 

USEPA (1992b) found that some organochlorine compounds (including dieldrin, oxychlordane, and DDE) were roughly similar in average 

concentrations in game-fish fillets and whole-fish samples of bottomfeeders, whereas other compounds (including chlordane, nonachlor, 

and heptachlor epoxide) had higher average concentrations in wholefish samples of bottom feeders than in game-fish fillets.

NAWQA results for whole fish, with these caveats considered, 

may be useful for screening-level assessment of streams for which 

there are no data specifically on edible tissue of fish commonly consumed in that area. If pesticide concentrations measured in a wholefish sample are less than a screening value for edible tissue, then 

residues in the edible portion of the fish are likely to be less than the 

screening value, suggesting low human-health concern. On the other 

hand, if a concentration in whole fish exceeds the screening value, 

the level in edible tissue may not exceed the value, but additional 

sampling and analysis of fillets for species that are commonly consumed may be warranted to determine whether or not the concentration in edible fish tissue exceeds the screening value.

The NAWQA analysis provides the following general perspective: 

Organochlorine concentrations measured by NAWQA in whole 

fish exceeded USEPA screening values most often in agricultural 

and urban streams (67 percent of sites), followed by streams 

draining areas of mixed land use (55 percent). 

Concentrations greater than screening values in agricultural 

streams were dominated by dieldrin, total DDT, and heptachlor 

epoxide, whereas these same compounds plus total chlordane 

accounted for most concentrations greater than screening values in urban streams.

 If people commonly consume fish from a stream where screening values were exceeded by NAWQA-measured concentrations in whole fish, and no prior monitoring of the commonly 

consumed fish has been done, then further investigation of 

organochlorine pesticide compounds in edible fish tissue may be 

warranted. 

•

•

•

Potential Effects of Fish Consumption on Human Health 

94 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 6

Ground Water

Concentrations of one or more pesticides 

were greater than human-health benchmarks in 

about 1 percent of sampled wells that are used for 

drinking water—including 17 of 2,356 domestic 

wells and 8 of 364 public-supply wells (table 

6–1). Many public-supply wells have some level 

of water treatment, which may or may not affect 

pesticide concentrations, whereas domestic wells 

generally have no treatment, so that samples 

usually represent the actual quality of water 

consumed. Shallow ground water in urban areas 

had the greatest proportion of sampled wells with 

concentrations of pesticides that were greater 

than one or more benchmarks, including 1 of 9 

public-supply wells, 3 of 17 domestic wells, and 

37 of 835 observation wells, for a total of about 

5 percent. About 1 percent of wells sampled in 

agricultural areas and about 1 percent of wells 

sampled in major aquifers had concentrations 

greater than one or more benchmarks. Wells with 

concentrations greater than benchmarks were 

widely scattered among 36 of the 187 groundwater studies across the Nation, including 11 of 

33 urban land-use studies, 10 of 53 agricultural 

land-use studies, and 15 of 92 major aquifer 

studies (fig. 6–2). Most of these studies with one 

or more benchmark exceedances had only 1 or 

2 wells with exceedances. All concentrations 

greater than benchmarks were accounted for by 

dieldrin (72 wells) and four other pesticides: 

dinoseb (4 wells), atrazine (4), lindane (2), and 

diazinon (1).

Of the pesticides analyzed by NAWQA, 

dieldrin is the primary pesticide identified by the 

screening-level assessment for further consideration regarding ground water. Of the 72 wells 

with dieldrin concentrations greater than its 

screening-level benchmark, 39 were shallow 

wells in urban areas (including 3 domestic wells 

and 1 public-supply well), 12 were shallow wells 

in agricultural areas (including 5 domestic wells), 

and 21 were wells in major aquifers (including 7 

domestic and 6 public-supply wells). Although 

aldrin (which transforms to dieldrin) and dieldrin are no longer used in the United States, the 

screening-level assessment indicates that some 

wells may still be affected by dieldrin from historical uses.

Table 6–1. Most wells sampled for agricultural and urban land-use studies were shallow observation wells that are not used for 

drinking water, but about 29 percent of wells sampled in agricultural areas were domestic wells. Most wells sampled for the major 

aquifer studies are used for drinking water; about 13 percent were public-supply wells, and 71 percent were domestic wells. Overall, 

about 1 percent of all domestic and public-supply wells had concentrations of a pesticide greater than a human-health benchmark.

Type of ground-water study

Public-supply wells Domestic wells Observation wells

Number 

sampled

Percentage of 

samples exceeding 

a benchmark

Number 

sampled

Percentage of 

samples exceeding 

a benchmark

Number 

sampled

Percentage of 

samples exceeding 

a benchmark

Agricultural land use 1 0.0 406 1.2 1,005 1.1

Urban land use 9 11 17 18 835 4.4

Major aquifers 354 2.0 1,933 0.5 453 2.0

Human-health benchmarks were seldom 

exceeded in domestic and public-supply wells.

Potential for Effects on Human Health, Aquatic Life, and Wildlife 95
Stream water

Bed sediment

Stream water

Bed sediment

Stream water

Bed sediment

Stream water

Bed sediment

0 25 50 75 100

Percentage of stream sites with

one or more pesticide compounds

exceeding an aquatic-life benchmark

Agricultural areas

Urban areas

Undeveloped areas

Mixed land uses

Concentrations greater than

aquatic-life benchmarks

Figure 6–4. Pesticides have the potential to 

adversely affect aquatic life in many streams, 

particularly in urban areas, as indicated by 

the relatively high proportions of sites with 

measured concentrations greater than aquatic-life 

benchmarks for both water and bed sediment.

of 65 streams with mixed-land-use watersheds 

had concentrations of at least one pesticide that 

exceeded one or more aquatic-life benchmarks 

during the selected year of sampling (fig. 6–4). 

One of 8 undeveloped streams that were sampled 

for pesticides in water had concentrations that 

were greater than an aquatic-life benchmark. 

Concentrations greater than benchmarks occurred 

throughout the study period. Agricultural streams 

had benchmark exceedances at 68 percent of sites 

sampled during 1993–1994, 43 percent during 

1995–1997, and 50 percent during 1998–2000. 

Urban streams had benchmark exceedances at 

90 percent of sites sampled during 1993–1994, 

100 percent during 1995–1997, and 64 percent 

during 1998–2000. Streams with mixed land uses 

in their watersheds had benchmark exceedances 

at 38 percent of sites sampled during 1993–1994, 

40 percent during 1995–1997, and 46 percent 

during 1998–2000.

Streams in which one or more pesticides 

exceeded an aquatic-life benchmark for water are 

distributed throughout the country in agricultural, 

urban, and mixed-land-use settings (fig. 6–5). 

Most concentrations that exceeded benchmarks, 

particularly by the greatest amounts, occurred 

during seasonal periods of high concentrations, 

as illustrated by results for diazinon in Arcade 

Creek, an urban stream in the Sacramento River 

Basin (fig. 6–6). The number, type, and degree 

of benchmark exceedances vary widely among 

sites indicated in figure 6–5 and meaningful 

generalizations are difficult. Some streams, 

such as Arcade Creek (fig. 6–6), exceeded one 

or more benchmarks by substantial margins for 

a sustained period during the year. Other sites 

briefly exceeded a benchmark for one pesticide. 

Of the 100 sites with one or more benchmark 

exceedances, 46 sites exeeded 1 benchmark to 

varying degrees and frequencies, and 30 sites 

exceeded 3 or more different benchmarks to 

varying degrees and frequencies. Because of 

this variability and the complexity of translating 

exceedances of screening-level benchmarks into 

specific potential for effects, the screening-level 

results, as noted earlier should be used as the 

starting point for further site-specific investigation. Streams in which concentrations did not 

exceed a benchmark included most undeveloped 

streams, plus streams in agricultural and mixedland-use settings in regions where pesticide use 

was low, such as the Yellowstone River Basin and 

the Ozark Plateaus.

Screening-Level Assessment for 

Aquatic Life in Streams 

The potential for pesticides to adversely 

affect aquatic life in streams was evaluated by 

comparing measured concentrations in water and 

bed sediment with their respective water-quality 

benchmarks. The benchmarks are described in 

the accompanying sidebars (p. 97 and 105) and 

benchmark values are listed in Appendix 3.

Water 

NAWQA findings indicate that pesticides 

detected in stream water, most of which were 

in use during the study period, had the potential 

to adversely affect aquatic life in many of the 

streams sampled. Of 186 stream sites sampled 

nationwide, 57 percent of 83 agricultural streams, 

83 percent of 30 urban streams, and 42 percent 

96 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Aquatic-Life Benchmarks for Pesticides 

in Water 

Benchmarks for assessing the potential for pesticides in stream 

water to adversely affect aquatic life were of two general types: (1) 

ambient water-quality criteria for the protection of aquatic life (AWQCAL), which were developed by USEPA’s Office of Water (OW), and (2) 

benchmarks derived from toxicity values obtained from registration and 

risk-assessment documents developed by USEPA’s Office of Pesticide 

Programs (OPP). Toxicity data from OPP documents were used to supplement OW criteria to expand the coverage of pesticides and to incorporate the most recent toxicity information used by USEPA. AWQC-AL are 

available for 7 of the 83 pesticides and degradates analyzed by NAWQA. 

One or more toxicity values from OPP documents are available for 60 of 

the 83 NAWQA analytes, including 5 of the 7 that have AWQC-AL. A total 

of 62 of the pesticide compounds analyzed in water by NAWQA have one 

or more aquatic-life benchmarks (Appendix 3A).

Ambient Water-Quality Criteria for Aquatic Organisms

USEPA’s OW derives both acute and chronic criteria, each of which 

specifies a threshold concentration for unacceptable potential for 

effects, an averaging period, and an acceptable frequency of exceedances. 

Acute AWQC-AL—The highest concentration of a chemical to which an 

aquatic community can be exposed briefly without resulting in an unacceptable effect. Except where a locally important species is very sensitive, aquatic organisms should not be unacceptably affected if the 1-hour 

average concentration does not exceed the acute criterion more than 

once every 3 years, on average. The intent is to protect 95 percent of a 

diverse group of organisms (USEPA, 2004d). 

Chronic AWQC-AL—The highest concentration of a chemical to which 

an aquatic community can be exposed indefinitely without resulting in 

an unacceptable effect. Except where a locally important species is very 

sensitive, aquatic organisms should not be unacceptably affected if the 

4-day average concentration does not exceed the chronic criterion more 

than once every 3 years, on average. The intent is to protect 95 percent 

of a diverse group of organisms (USEPA, 2004d). 

Toxicity Values from Risk Assessments

Seven types of aquatic toxicity values were compiled from OPP’s 

registration and risk-assessment documents. The OPP toxicity values are 

for specific types of organisms. Acute and chronic values were compiled 

for fish and invertebrates, and acute values for vascular and nonvascular 

plants. A value for aquatic-community effects was available only for atrazine. The types and amounts of toxicity data available for different pesticides were highly variable. USEPA estimates the toxicity or hazard of a 

pesticide by selecting the most sensitive endpoints from multiple acute 

and chronic laboratory and field studies. For many pesticides, USEPA has 

completed a screening-level ecological risk assessment, which includes 

acute and chronic assessments for both fish and invertebrates. For some 

pesticides, acute assessments have also been completed for nontarget 

aquatic plants. NAWQA derived benchmarks from OPP toxicity values, 

generally following OPP procedures (USEPA, 2005h).

In recent years, USEPA has developed methods for conducting 

refined risk assessments, in which probabilistic tools and methods are 

incorporated to predict the magnitude of the expected impact of pesticide use on nontarget organisms, as well as the uncertainty and variability involved in these estimates. The screening-level benchmarks used in 

NAWQA analysis and summarized below were derived from the toxicity 

values reported in USEPA registration and risk-assessment documents. 

In the few cases where refined assessments were available, these 

were given preference. In deriving a benchmark for a given type of 

organism (such as fish) and a given exposure duration (acute or chronic), 

the lowest of the available toxicity values was selected for each benchmark, unless a preferred toxicity value was specified in a refined risk 

assessment—in which case that preferred toxicity value was used 

instead. For two of the benchmarks—acute-fish and acute-invertebrates—the selected toxicity values were multiplied by the USEPA level 

of concern (LOC) of 0.5, so that the benchmark for NAWQA screening 

corresponds to the acute risk level defined by USEPA (2005h). 

Six benchmarks were based directly on toxicity endpoints used in 

OPP screening-level assessments (USEPA, 2005i):

Acute fish—The lowest tested 50-percent lethal concentration (LC50) for 

acute (typically 96-hour) toxicity tests with freshwater fish, multiplied by 

the LOC of 0.5.

Acute invertebrate—The lowest tested LC50 or 50-percent effect concentration (EC50) for acute (typically 48 or 96-hour) toxicity tests with freshwater invertebrates, multiplied by the LOC of 0.5.

Acute vascular plant—The lowest tested EC50 for freshwater vascular 

plants in acute toxicity tests (typically < 10 days).

Acute nonvascular plant—The lowest tested EC50 for freshwater nonvascular plants (algae) in acute toxicity tests (typically < 10 days). 

Chronic fish—The lowest no-observed-adverse-effects concentration 

(NOAEC), or the lowest-observed-adverse-effects concentration (LOAEC) 

if a NOAEC is not available, for freshwater fish in early lifestage or full 

life-cycle tests. 

Chronic invertebrate—The lowest NOAEC, or LOAEC if a NOAEC is not 

available, for freshwater invertebrates in life-cycle tests.

One additional benchmark, a benchmark for aquatic-community effects, 

was derived from the refined risk assessment for atrazine. This endpoint 

for atrazine incorporates community-level effects on aquatic plants and 

indirect effects on fish and aquatic invertebrates that could result from 

disturbance of the plant community (USEPA, 2003b).

Application of Aquatic-Life Benchmarks for Water

Acute AWQC-AL values and all acute benchmarks were compared 

with each measured concentration for the most complete year of data 

for each NAWQA stream site. Chronic AWQC-AL values were compared 

with 4-day moving average concentrations. This approach matches the 

time periods in the definitions of acute and chronic AWQC-AL, which are 

1-hour average and 4-day average concentrations, respectively (Stephan 

and others, 1985). Chronic benchmarks for invertebrates were compared 

with 21-day moving averages, and chronic benchmarks for fish and the 

aquatic-community benchmark for atrazine were compared with 60-day 

moving averages. These time periods are those used or recommended by 

USEPA in OPP risk assessments (USEPA, 2003b; USEPA, 2005g). Moving 

average concentrations for 4-, 21-, and 60-day periods were computed 

for each day of the year for each stream site from hourly concentration 

estimates determined by straight-line interpolation between samples. 

This method was tested using data on pesticide concentrations in Ohio 

streams studied by Richards and Baker (1993) and Richards and others 

(1996), using an approach similar to that used by Crawford (2004). Results 

indicate that all three averages, but particularly the 4-day averages, 

are consistently underestimated when computed from data collected 

at frequencies similar to the NAWQA sampling design (indicating a tendency to also underestimate the potential for toxicity to aquatic life in this 

respect). 

Potential for Effects on Human Health, Aquatic Life, and Wildlife 97

Chapter 6
Land use

Agricultural

Urban

Undeveloped

Mixed

One or more

benchmarks exceeded

Stream sites

No benchmark

exceeded

Stream Water

Sites with concentrations greater than an aquatic-life benchmark

Bed Sediment

##

##

#

##

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## ####

# ### #########

# ###

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##

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# #

### # # # # #### ## # # # ###

## # ## ### # #####

#

#

#

#

#

######

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### ##

###### #

##########

#

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#

#

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######

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####

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## # ## # ###

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# # ## ##

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########

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604d

#

#

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#

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# ### #####

#

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#

Figure 6–5. Aquatic-life benchmarks for pesticides in water and bed sediment were exceeded by concentrations 

measured in many agricultural, urban, and mixed-land-use streams throughout the Nation. 

Diazinon

Water-quality benchmark 

for acute effects on 

invertebrates (0.1μg/L)

6_05arcade

Diazinon in Arcade Creek, California 

0

0.5

1.5

1

1996 1997 1998

Concentration, in micrograms per liter

NovemberDecember

January

February

March

April

May

June

July

August

September

October

NovemberDecember

January

February

March

April

Figure 6–6. Concentrations of diazinon 

in Arcade Creek, an urban creek in 

Sacramento, California (Sacramento 

River Basin), exceeded the aquaticlife benchmark for acute effects on 

invertebrates (0.10 μg/L) by the greatest 

amounts during seasonal pulses of high 

concentrations in the winter and spring. 

(Modified from Domagalski and others, 

2000.)

98 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 6

Percentage of stream sites

exceeding one or more benchmarks

Percentage of stream sites

exceeding one or more benchmarks

Diazinon

Chlorpyrifos

Malathion

Parathion-methyl

Dieldrin

p,p-DDE

Parathion

Diuron

Carbaryl

Azinphos-methyl

Chlorpyrifos

Azinphos-methyl

Atrazine

p,p-DDE

Alachlor

Diazinon

Malathion

Parathion-methyl

Carbofuran

Disulfoton

Diuron

Methomyl

Thiobencarb

Carbaryl

Dieldrin

Molinate

Parathion

Phorate

Propargite

Terbufos

Agricultural streams

Pesticides with concentrations

greater than an aquatic-life

benchmark

Urban streams

0 25 50 75 100

0 25 50 75 100

Figure 6–7. Contributions of individual pesticides 

to exceedances of aquatic-life benchmarks for 

water show the significance of insecticides in 

urban streams, particularly diazinon, chlorpyrifos, 

and malathion during the 1992–2001 study period. 

In agricultural streams, most exceedances of 

benchmarks were by chlorpyrifos, azinphos-methyl, 

atrazine, p,p’-DDE, and alachlor. Water-quality 

benchmarks are provided in Appendix 3A. 

The insecticides diazinon, chlorpyrifos, and 

malathion accounted for most concentrations 

that were greater than aquatic-life benchmarks in 

water from urban streams, whereas chlorpyrifos, 

azinphos-methyl, atrazine, p,p'-DDE, and alachlor 

accounted for most concentrations greater than 

benchmarks in water from agricultural streams 

(fig. 6–7). Streams draining watersheds with 

mixed land uses reflected a combination of urban 

and agricultural influences. Generally, the types of 

benchmarks most frequently exceeded by the herbicides atrazine and alachlor were those for acute 

effects on either vascular or nonvascular plants, 

whereas the insecticides diazinon, chlorpyrifos, 

malathion, azinphos-methyl, and carbaryl most 

frequently exceeded acute or chronic benchmarks 

for invertebrates or benchmarks based on chronic 

ambient water-quality criteria.

The geographic distributions of benchmark 

exceedances for atrazine (fig. 6–8), diazinon 

(fig. 6–9), and chlorpyrifos (fig. 6–10) illustrate 

the varying distributions and types of potential 

effects on aquatic life. Concentrations of atrazine 

were greater than one or more aquatic-life benchmarks in 18 percent of agricultural streams, but 

in only one stream with a predominantly nonagricultural watershed. As discussed in Chapter 4, 

concentrations of atrazine in agricultural streams 

matched the geographic distribution of corn cultivation, where applications are greatest (fig. 4–9). 

As noted above, the atrazine benchmarks most 

frequently exceeded were the acute benchmarks 

for vascular and nonvascular plants, although the 

benchmark for aquatic community effects and the 

chronic benchmark for invertebrates also were 

exceeded at about 35 and 12 percent, respectively, of the sites where one or more atrazine 

benchmarks were exceeded (fig. 6–8). 

Diazinon concentrations were greater than 

one or more aquatic-life benchmarks in 73 

percent of the urban streams that were sampled, 

compared with 37 percent for chlorpyrifos 

(fig. 6–7). The urban stream sites where diazinon 

exceeded a benchmark were distributed throughout the country (fig. 6–9). Benchmarks for both 

of these insecticides were exceeded in smaller 

proportions of agricultural streams, although 

chlorpyrifos exceeded one or more of its benchmarks in 21 percent of the agricultural streams. 

The highest concentrations of chlorpyrifos in 

agricultural streams, as discussed in Chapter 4, 

were in streams within the corn-growing areas 

of the central United States; in the lower Mississippi River Basin, where both corn and cotton are 

grown; and in streams draining agricultural areas 

in the West, where fruits, nuts, and vegetables are 

grown.

The diazinon benchmarks most frequently 

exceeded (fig. 6–9) were the acute and chronic 

benchmarks for invertebrates reported by USEPA 

(USEPA, 2004e). As shown in figure 6–10, 

the chlorpyrifos benchmarks most frequently 

exceeded were the acute and chronic benchmarks 

for invertebrates and also the acute and chronic 

ambient aquatic-life criteria (Appendix 3A). 

While none currently exists, USEPA is drafting 

ambient aquatic-life criteria for diazinon. During 

Potential for Effects on Human Health, Aquatic Life, and Wildlife 99
100    Pesticides in the Nation’s Streams and Ground Water, 1992–2001

Estimated 1997 agricultural use intensity,

in pounds per square mile per year

Watershed

land use

Agricultural

Urban

Mixed

One or more

benchmarks exceeded

Stream sites

No benchmark

exceeded

EXPLANATION

Benchmarks exceeded at the 16 sites 

where exceedances occurred

< 0.09 

0.09 – 4.5

> 4.5 – 45

> 45

Atrazine concentrations in stream water compared with aquatic-life benchmarks

0

20

40

60

80

100

Exceedance of individual benchmarks,

as a percentage of all sites with one 

or more exceedances

Fish: acuteFish: chronicInvertebrates: acute

Invertebrates: chronicVascular plants: acuteNonvascular plants: acuteAquatic-life criterion: acuteAquatic-life criterion: chronicAquatic community effects

Atrazine benchmarks for aquatic life

no exceedancesno exceedancesno exceedancesno benchmarkno benchmark

#

#

## #

#

##

#

# ## ##

# ## ### ## # #

## # #

# ##### #

# ## ##

## ######

# #

# # # #

### #

# # # #

#

#

#

Agricultural Mixed land use

Urban

Figure 6–8.  Streams in which atrazine concentrations were greater than at least one of its aquatic-life benchmarks were

predominantly agricultural streams in areas where applications were greatest. The aquatic-life benchmarks most frequently

exceeded by atrazine concentrations were those for vascular and nonvascular plants. Water-quality benchmarks are provided

in Appendix 3A.
Potential for Effects on Human Health, Aquatic Life, and Wildlife 101

Chapter 6

Figure 6–9.  Most streams in which diazinon concentrations were greater than at least one aquatic-life benchmark

were urban streams, but concentrations in some agricultural streams in areas where applications were greatest

also exceeded a benchmark. The aquatic-life benchmarks most frequently exceeded by diazinon were those for

invertebrates. Water-quality benchmarks are provided in Appendix 3A.

Estimated 1997 agricultural use intensity,

in pounds per square mile per year Watershed

land use

Agricultural

Urban

Mixed

One or more

benchmarks exceeded

Stream sites

No benchmark

exceeded

EXPLANATION

Diazinon concentrations in stream water compared with aquatic-life benchmarks

Agricultural Mixed land use

Urban

Diazinon benchmarks for aquatic life

Fish: acuteFish: chronicInvertebrates: acuteInvertebrates: chronicVascular plants: acuteNonvascular plants: acuteAquatic-life criterion: acuteAquatic-life criterion: chronic

0

20

40

60

80

100

Exceedance of individual benchmarks, 

as a percentage of all sites with one 

or more exceedances

no exceedancesno benchmarkno exceedancesno benchmarkno benchmark

#

#

# #

##

##

## ##

# ## ### ## # #

# # ##

# # #

# # ## ###

# #

# # # #

### #

# # # #

#

#

#

#

# ###

# # ###

#

#

Benchmarks exceeded at the 44

sites where exceedances occurred

< 0.09 

0.09 – 4.5

> 4.5 – 45

> 45
102    Pesticides in the Nation’s Streams and Ground Water, 1992–2001

Estimated 1997 agricultural use intensity,

in pounds per square mile per year Watershed

land use

Agricultural

Urban

Mixed

One or more

benchmarks exceeded

Stream sites

No benchmark

exceeded

EXPLANATION

Chlorpyrifos concentrations in stream water compared with aquatic-life benchmarks

Agricultural Mixed land use

Urban

0

20

40

60

80

100

Chlorpyrifos benchmarks for aquatic life

no exceedancesno exceedancesno benchmarkno exceedances

#

#

## #

##

##

# ## #

# ## ### ## #

# # # ## #

# #

# ## #### # # # #

# #

# # # #

### #

# # # #

#

# #

#

# # #

# #

#

Fish: acuteFish: chronicInvertebrates: acuteInvertebrates: chronic

Vascular plants: acuteNonvascular plants: acuteAquatic-life criterion: acuteAquatic-life criterion: chronic

Exceedance of individual benchmarks,

as a percentage of all sites with one 

or more exceedances

Benchmarks exceeded at the 37

sites where exceedances occurred

< 0.09 

0.09 – 4.5

> 4.5 – 45

> 45

Figure 6–10.  Most streams in which chlorpyrifos concentrations were greater than at least one aquatic-life benchmark were

agricultural streams in areas where applications were greatest, or urban streams. The aquatic-life benchmark most frequently

exceeded by chlorpyrifos was the USEPA chronic aquatic-life criterion. Chlorpyrifos concentrations also frequently exceeded

acute and chronic benchmarks for invertebrates at the same sites where the chronic aquatic-life criterion was exceeded.

Water-quality benchmarks are provided in Appendix 3A.
Chapter 6

this development process, and in response to 

USEPA’s diazinon risk assessment, public comment noted an atypical distribution of the acute 

toxicity data for invertebrates. If data from the 

second most sensitive study were used (USEPA, 

2000b), rather than the most sensitive study, then 

the calculated acute invertebrate benchmark for 

diazinon would change from its original 0.1 μg/L 

to a value of 0.4 μg/L. The result of using a 

benchmark of 0.4 μg/L would be a reduction in 

the proportions of sites with diazinon exceedances from 73 to 40 percent for urban streams 

and from 8 to 6 percent for agricultural streams.

Overall, the screening-level assessment for 

potential effects of pesticides in stream water on 

aquatic life indicates that 56 percent of the 178 

sampled streams that have watersheds dominated 

by urban, agricultural, or mixed land uses had 

concentrations of one or more pesticides that 

exceeded an aquatic-life benchmark during the 

study period. Pesticide use and occurrence were 

not constant during 1992–2001, however, and 

NAWQA data can be used, as for human-health 

benchmarks, to characterize changes that may 

have occurred for some pesticides in the land-use 

settings for which there are adequate data.

As noted for analysis of human-health 

benchmarks, there are sufficient NAWQA data 

for limited analysis of changes over time in 

benchmark exceedances for urban streams and 

for agricultural streams in the corn-and-soybeans 

crop setting. When grouped by sampling period, the percentages of urban stream sites that 

had concentrations of diazinon, chlorpyrifos, or 

malathion that exceeded a benchmark were lowest for urban sites sampled during the last part of 

the study (fig. 6–11). Observations about changes 

shown in figures 6–11 and 6–12, however, are 

preliminary because they are based on different 

groups of sites for each sampling period and siteto-site variability in conditions may distort actual 

trends. Although there are no consistent data 

available on the trends in the urban use of these 

pesticides during the study period, these results 

indicate the possibility that some reductions in 

urban use may have occurred. As mentioned earlier and discussed in Chapter 8, nonagricultural 

uses and some agricultural uses of diazinon and 

chlorpyrifos have declined since 2001 because of 

use restrictions initiated by USEPA. If concentrations of these insecticides are, in fact, declining in urban streams, the potential for effects 

on aquatic life in urban streams likely will also 

decline if their uses are replaced with pesticides 

that reach streams in less toxic amounts (or with 

alternative approaches to pest control).

In agricultural streams, most exceedances 

of aquatic-life benchmarks were by chlorpyrifos, 

azinphos-methyl, atrazine, p,p'-DDE, and alachlor (fig. 6–7). The greatest potential for effects 

on aquatic life was generally in areas where one 

or more of these pesticides were intensively 

used, or in the case of p,p'-DDE, where its parent 

compounds were intensively used in the past. 

For the purpose of characterizing changes over 

time in benchmark exceedances, there were sufficient agricultural stream sites with sampling 

years distributed throughout the study period 

only for streams in the corn-and-soybeans crop 

setting (fig. 4–6). This agricultural setting had the 

highest use during the study period of chlorpyrifos, atrazine, and alachlor. The changes in the 

percentages of stream sites in this setting that 

had concentrations exceeding benchmarks were 

different for the three pesticides during the study 

period (fig. 6–12). There was no clear trend for 

chlorpyrifos, an increasing number of exceedances for atrazine, and a decrease in exceedances 

for alachlor (with none during 1998–2000). Data 

on the agricultural use of these three pesticides 

from 1992 to 2001 in the Corn Belt show that 

these changes over time in benchmark exceedances are consistent with changes in their use 

(fig. 6–12). 

Figure 6–11. The percentages of urban streams 

that had exceedances of aquatic-life benchmarks 

for chlorpyrifos, malathion, and diazinon were 

lowest for each insecticide during the last sampling 

period. Sites were grouped according to the year of 

sampling. The 1993–1994 sampling period included 

10 sites, the 1995–1997 period included 9 sites, and 

the 1998–2000 period included 11 sites.

1993-19941995-19971998-20001993-19941995-19971998-20001993-19941995-19971998-2000

Stream sites with exceedances 

of one or more benchmarks, 

as a percentage of sites sampled 

during each sampling period 

Chlorpyrifos Malathion Diazinon

0

100

75

50

25

Sampling period

Trends in aquatic-life benchmark exceedances for

chlorpyrifos, malathion, and diazinon in urban streams

Potential for Effects on Human Health, Aquatic Life, and Wildlife 103
0 25 50 75 100

0 25 50 75 100

Percentage of stream sites exceeding

the benchmark

Percentage of stream sites exceeding

the benchmark

DDT compounds (TEC)

Total chlordane (TEC)

Dieldrin (TEC)

Heptachlor epoxide (TEC)

Methoxychlor (ESB)

Endrin (TEC)

DDT compounds (TEC)

Dieldrin (TEC)

Total chlordane (TEC)

Methoxychlor (ESB)

Endrin (TEC)

Heptachlor epoxide (TEC)

Toxaphene (ESB)

Agricultural streams 

Organochlorine compounds with concentrations 

in bed sediment greater than an aquatic-life

benchmark

Urban streams 

Figure 6–13. Contributions of individual pesticide 

compounds and groups to exceedances of aquatic-life 

benchmarks for bed sediment show the importance 

of historically used insecticides in urban streams, 

particularly DDT compounds, chlordane, and dieldrin. 

In agricultural streams, DDT compounds and dieldrin 

accounted for most exceedances of benchmarks. The 

type of benchmark is listed after each compound name 

as ESB for equilibrium partitioning sediment benchmark, 

or as TEC for threshold effect concentration. Waterquality benchmarks are provided in Appendix 3B. 

Figure 6–12. Changes over time in the percentage of 

agricultural stream sites in the corn-and-soybeans crop 

setting that had exceedances of aquatic-life benchmarks 

for chlorpyrifos, atrazine, and alachlor generally followed 

trends in use. Sites were grouped according to the year of 

sampling. The 1993–1994 sampling period included 10 sites, 

the 1995–1997 period included 19 sites, and the 1998–2000 

period included 6 sites. 1993-1994 1995-1997 1998-2000

1993-19941995-19971998-20001993-19941995-19971998-2000

Sampling period

0.1

0

0.2

0.3

0.4

1992 1993 1994 1995 1996 1997 1998 1999 2000 2001

25

0

50

75

100

Use intensity, in pounds per acre 

of total agricultural land

Chlorpyrifos Atrazine Alachlor

Chlorpyrifos

Atrazine

Alachlor

Trends in use and aquatic-life benchmark 

exceedances for chlorpyrifos, atrazine, and alachlor 

Agricultural use of chlorpyrifos, atrazine, and alachlor 

(IA, IL, IN, MN, NE, OH)

Benchmark exceedances for agricultural streams in corn-andsoybeans crop setting

Stream sites with exceedances 

of a benchmark, as a percentage of sites 

sampled during each sampling period

no exceedances

sites, total chlordane at 57 percent of sites, and 

dieldrin at 26 percent of sites (fig. 6–13). In agricultural streams, compounds in the DDT group 

exceeded benchmarks at 28 percent of sites and 

dieldrin at 8 percent of sites. 

The geographic distributions of concentrations that were greater than benchmarks are different for DDT compounds (fig. 6–14) compared 

with dieldrin (fig. 6–15), following their historical use patterns. Concentrations of one or more 

DDT compounds were greater than benchmarks 

for aquatic life in 58 percent of urban streams 

and about 28 percent of agricultural and mixedland-use streams that were sampled. As discussed 

in Chapter 4, historical use of DDT for agriculture was highest in the Southeast, where cotton, 

tobacco, and peanuts were grown, and in a number of areas of the Nation where orchard crops, 

potatoes, vegetables, or specialty crops were 

grown. Dieldrin concentrations did not exceed 

its aquatic-life benchmark as frequently as DDT 

compounds, with 26 percent of urban streams 

Bed Sediment

Concentrations of organochlorine pesticide 

compounds measured in bed sediment were 

greater than one or more aquatic-life benchmarks 

at 70 percent of urban sites, 31 percent of agricultural sites, 36 percent of mixed-land-use sites 

and 8 percent of undeveloped sites (fig. 6–4). The 

geographic distribution of sites where aquatic-life 

benchmarks for bed sediment were exceeded is 

similar to findings for water in many respects, 

including urban streams distributed throughout 

the country and many agricultural and mixedland-use streams in the Southeast, East, and 

irrigated areas of the West (fig. 6–5). 

In urban streams, concentrations of DDT or 

one or more of its degradates or by-products were 

greater than benchmarks at 58 percent of sampled 

104 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 6

percent), and about one-third of sampled streams 

with watersheds dominated by agricultural or 

mixed land uses, had concentrations of organochlorine compounds that exceeded one or more 

aquatic-life benchmarks during the study period. 

Although DDT, aldrin, dieldrin, and chlordane 

are no longer used in the United States, the 

screening-level assessment indicates that these 

compounds and their degradates continue to be 

present at levels in bed sediment that may have 

adverse effects on aquatic life in some streams. 

and 8 percent of agricultural and mixed-land-use 

streams having concentrations greater than the 

benchmark. For dieldrin, a cluster of agricultural 

sites with concentrations greater than the benchmark is located in the Corn Belt, where use of 

aldrin and dieldrin on corn was most intensive. In 

urban areas, these pesticides were used for such 

purposes as mosquito and termite control.

The screening-level assessment for organochlorine compounds in bed sediment indicates 

that most urban streams sampled by NAWQA (70 

Aquatic-Life Benchmarks for 

Organochlorine Compounds in Bed 

Sediment 

Benchmarks for assessing the potential for organochlorine 

pesticides compounds in bed sediment to adversely affect aquatic 

life were selected from consensus-based sediment-quality 

guidelines developed for sediment-dwelling aquatic organisms 

(MacDonald and others, 2000). These benchmarks are available 

for 6 of the 16 individual organochlorine pesticide compounds and 

compound groups (such as total chlordane) measured in sediment, including all of the most commonly detected ones. Threshold 

effect concentrations (TEC), which are concentrations below 

which harmful effects on sediment-dwelling organisms are not 

expected, were used as the primary screening-level benchmarks. 

In NAWQA’s analysis, the TEC benchmarks were supplemented 

by USEPA equilibrium partitioning sediment benchmarks (ESB), 

which are available for 6 of the 16 organochlorine pesticide 

compounds and groups measured (USEPA, 2003c,d,e). Although 

ESBs are not available for some of the most commonly detected 

pesticides in sediment (DDT and chlordane), the 6 compounds 

with ESBs include 3 pesticides that do not have TEC benchmarks—toxaphene, methoxychlor, and endosulfan (Appendix 3B). 

Therefore, sediment benchmarks are available for a total of 9 of 

the 16 organochlorine pesticides or pesticide groups analyzed by 

NAWQA. 

The two types of sediment benchmarks are quite different 

from one another. The TECs are empirically derived and are effective predictors of toxicity (or nontoxicity) in field-collected sediment, but they cannot be used to infer cause and effect related 

to individual contaminants. The TEC was selected as the primary 

benchmark because it meets the objectives of a screening-level 

assessment. The ESB is mechanistically based and is not designed 

to predict toxicity in field-collected sediment that contains multiple 

contaminants. A concentration greater than an ESB indicates a 

high likelihood of toxicity resulting from the specific contaminant. 

ESBs were used to provide some information on potential toxicity 

for pesticides that do not have TEC benchmarks.

Consensus-based threshold effect concentration (TEC)—The 

concentration of sediment-associated contaminants below which 

adverse effects on sediment-dwelling organisms are not expected 

to occur. The consensus-based TEC benchmarks are empirically 

based and indicate the likelihood that field-collected samples containing a given pesticide concentration will be toxic or nontoxic, 

but they do not necessarily indicate cause-and-effect. The particular pesticide upon which the benchmark is based is not necessarily the source of the toxicity because sediment may contain 

multiple contaminants. Validation data showed that 15–29 percent 

of sediment samples, depending on the pesticide, had measurable 

toxicity at organochlorine pesticide concentrations below their 

respective TECs (MacDonald and others, 2000). The incidence of 

toxicity above the TEC was consistently higher, with 40 percent of 

samples for one pesticide (endrin), and 70–100 percent for the rest, 

showing measurable toxicity above their respective TECs. 

Equilibrium partitioning sediment benchmark (ESB)—The concentration of a chemical in sediment that USEPA expects will not 

adversely affect most benthic organisms. ESBs are mechanistic 

benchmarks based on the equilibrium partitioning model, which 

assumes that the toxicity of an organic contaminant in sediment 

is causally related to bioavailability and that bioavailability is 

controlled by contaminant sorption to sediment organic carbon. 

ESBs further assume that the contaminant is in equilibrium with 

sediment particles and sediment pore water. In the natural environment, including areas with highly erosional or depositional 

bed sediment, contaminants may not attain equilibrium. Each ESB 

is designed to predict toxicity caused by a specific contaminant 

(or group) only, and it is not expected to correctly predict toxicity 

when other contaminants are present in toxic amounts, such as 

may occur in field-collected samples containing contaminant mixtures. Thus, when a contaminant concentration exceeds its ESB 

in field-collected sediment, the sediment is predicted to be toxic 

because of the presence of that contaminant.

Application of Aquatic-Life Benchmarks for Bed 

Sediment

Aquatic-life benchmarks for sediment, both TECs and ESBs, 

were compared with pesticide concentrations measured by 

NAWQA in composite bed-sediment samples collected from depositional areas in streams (one sample per site). TECs, which are 

expressed on a total sediment basis, were compared directly with 

NAWQA-measured pesticide concentrations in sediment. Because 

ESBs are in units of micrograms of contaminant per gram of sediment organic carbon, NAWQA-measured pesticide concentrations 

(micrograms of contaminant per kilogram of total sediment) were 

first divided by the measured organic carbon content (grams of 

organic carbon per kilogram of total sediment) of the sediment 

sample, before comparison with ESBs. 

Potential for Effects on Human Health, Aquatic Life, and Wildlife 105
106    Pesticides in the Nation’s Streams and Ground Water, 1992–2001

Estimated historical agricultural use intensity,

in pounds per square mile per year

Agricultural

Urban

Undeveloped

Mixed

EXPLANATION

Watershed

land use

One or more

benchmarks exceeded

Stream sites

No benchmark

exceeded

< 0.09 

0.09 – 4.5

> 4.5 – 45

> 45

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Concentrations of DDT compounds in bed sediment compared with aquatic-life benchmarks

Urban Undeveloped

Agricultural Mixed land use

Figure 6–14.   Streams in which concentrations of one or more DDT compounds in bed sediment exceeded an

aquatic-life benchmark were predominantly urban streams, or agricultural and mixed-land-use streams in areas where

historical use of DDT plus DDD was greatest. Water-quality benchmarks are provided in Appendix 3B.
Potential for Effects on Human Health, Aquatic Life, and Wildlife 107

Chapter 6

Figure 6–15.   Streams in which dieldrin concentrations in bed sediment exceeded its aquatic-life benchmark were

predominantly urban streams or agricultural and mixed-land-use streams in areas where historical use of aldrin plus

dieldrin was greatest. Water-quality benchmarks are provided in Appendix 3B.

Concentrations of dieldrin in bed sediment compared with aquatic-life benchmarks

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Estimated historical agricultural use intensity,

in pounds per square mile per year

Agricultural

Urban

Undeveloped

Mixed

EXPLANATION

Watershed

land use

One or more

benchmarks exceeded

Stream sites

No benchmark

exceeded

< 0.09 

0.09 – 4.5

> 4.5 – 45

> 45

Undeveloped

Agricultural Mixed land use

Urban
Screening-Level Assessment for 

Fish-Eating Wildlife

NAWQA data on pesticides in whole fish 

were compared with both the low and high values 

of the range in available benchmarks, because 

there is no consensus on a national-scale suite of 

wildlife benchmarks (see accompanying sidebar, 

p. 109). Comparisons of measured concentrations 

of organochlorine pesticide compounds in wholefish tissue with wildlife benchmarks indicate 

a correspondingly wide range of potential for 

effects, depending on whether the low or high 

benchmark values are used (fig. 6–16). The high 

benchmarks for fish tissue were exceeded most 

frequently in streams in the populous Northeast; in high-use agricultural areas in the upper 

and lower Mississippi River Basin; in high-use 

irrigated agricultural areas of the West, such 

as eastern Washington and the Central Valley 

of California; and in urban streams distributed 

throughout the country (fig. 6–17). Few fish 

samples were analyzed in the Southeast. The low 

(more protective) benchmarks generally show an 

expanded proportion of sites in the same regions 

and land uses. 

0 25

Percentage of stream sites with one or more

organochlorine compounds exceeding a

wildlife benchmark

50 75 100

High benchmark

Low benchmark

Agricultural areas

Concentrations in whole-fish tissue

greater than wildlife benchmarks

Urban areas

Undeveloped areas

Mixed land uses

Figure 6–16. Wildlife 

benchmarks for 

concentrations of 

organochlorine 

pesticides in fish 

tissue were often 

exceeded, but the 

range of results for high 

and low benchmark 

values indicates that 

there is considerable 

uncertainty in 

wildlife benchmark 

values. Water-quality 

benchmarks are 

provided in Appendix 

3B. 

Agricultural

Urban

Undeveloped

Mixed

Watershed

land use

One or more

benchmarks exceeded

Stream sites

No benchmark

exceeded

Fish Tissue (high benchmarks for wildlife)

Sites with concentrations greater than 

a wildlife benchmark

Fish Tissue (low benchmarks for wildlife)

##

#####

##

##

####

# #### ##

###

## ##

# #

#

# # # # ## ## # # ###

#

# ###

#

## ###

# # ### ## #

## # ## # ##

# #

#

#######

#

# ##

## #########

#####

#####

##

# ##

# #

## # #

##

###

########## # ## ### ###

# # ###

# # # # #

##

#

# #

#

# # # ###

# # ##

# ## # ##

#

####

#

####

#

###

##

#

# #

###

#

#

##

##

# # #

#

#

#

##

#####

#

##

#

##

### ##

# ###

### ##

#

#

# #

#

#

#

## #

# ## # ## # ###

#

# ##

# ###

## # #

#### #

##

##

# ### #

# ###

######

#

# ##

## #########

##########

#

###### # # ## #

# #

#

# #

##

#

############### # ## ## ###

# # ###

## # # # #

Figure 6–17. Wildlife benchmarks were exceeded 

by organochlorine pesticide compounds in whole 

fish most frequently in urban and mixed-land-use 

streams in the populous Northeast, in agricultural 

streams in areas with high historical use, and 

in urban streams distributed throughout the 

country. Water-quality benchmarks are provided in 

Appendix 3B. 

108 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Benchmarks for assessing the potential for organochlorine 

pesticide compounds in fish tissue to adversely affect wildlife that 

consume either fish or other fish-eating wildlife were selected from 

several sources (Appendix 3B). USEPA is developing tissue-based 

criteria for bioaccumulative contaminants, but the process is not 

complete (USEPA, 2005i).

Currently, there is no broad consensus on a single system of 

national-scale, fish-tissue benchmarks for wildlife. Relatively few 

tissue-based wildlife benchmarks are available, some of which 

were developed for State or regional applications. Most available 

benchmarks have, however, been derived using similar methodologies (based on the same USEPA methodology for using laboratory 

animal test data to develop human-health benchmarks). First, a noobserved-adverse-effects level (NOAEL) for wildlife is estimated from 

the NOAEL for the most sensitive test species. Then the concentration 

of a contaminant in food that would result in a dose equivalent to the 

NOAEL (assuming no exposure through other environmental media) 

is calculated from estimates of the food consumption rate and body 

weight for multiple representative wildlife species. Calculations usually are done for both mammalian and avian species, and the lowest 

is commonly selected as a screening-level benchmark. Benchmark 

values from different sources vary considerably for a given compound, despite similar methodologies. The extreme case is total DDT, 

for which tissue-based wildlife benchmarks range from 6 to 200 μg/kg 

wet weight. Different values for a particular pesticide may result from 

the use of different test species in toxicity tests, the use of different 

uncertainty factors to account for interspecies differences, and differences in the duration of exposure or test endpoints measured. 

In addition, results may be extrapolated to different representative 

wildlife species, which typically are selected to reflect the geographic 

location and objectives of the program or organization setting the 

benchmarks.

Because of the lack of consensus on tissue-based benchmarks 

for protection of wildlife, whole-fish concentrations measured by 

NAWQA were compared with a range of available benchmark values 

for each compound. First, systematically derived wildlife benchmarks were compiled, resulting in four sets of wildlife benchmarks 

(described below). Second, the lowest and highest benchmark values 

for each organochlorine pesticide or group were selected and used 

in two separate analyses of NAWQA fish data. Each wildlife benchmark used in this report represents the concentration of a pesticide 

or group in fish, below which adverse effects on fish-eating wildlife 

are not expected to occur (100 percent of exposure to the pesticide 

is assumed to be from consumption of fish). One or more fish-eating 

wildlife benchmarks were available for 10 of the 12 organochlorine 

pesticides and groups measured by NAWQA in fish tissue. 

NOAEL-based toxicological benchmark for fish-eating wildlife—

This benchmark is the NOAEL-equivalent concentration in food 

derived for the most sensitive fish-eating wildlife species for which 

data are available. NOAEL-equivalent concentrations in food were 

derived for a variety of wildlife species by Sample and others (1996) 

for the Department of Energy, Oak Ridge National Laboratory, for use 

in ecological risk assessments at waste sites. Endpoints such as 

reproductive and developmental toxicity and reduced survival were 

used whenever possible, but for some contaminants, data were limited and other endpoints (such as organ-specific toxic effects) were 

used. The representative wildlife species used by Sample and others 

(1996) represent a wide range of diets and body weights and have 

wide geographic distributions within the United States. These include 

several fish-eating species: mink, river otter, belted kingfisher, osprey, 

and great blue heron. For this report, the lowest value was selected 

from the available NOAEL-equivalent concentrations in food that were 

derived for fish-eating species and used as the benchmark for each 

compound. These benchmarks are available for 8 of the 12 organochlorine pesticides and pesticide groups measured by NAWQA in fish.

Canadian Tissue Residue Guideline (TRG)—This benchmark is 

designed to protect all life stages of all wildlife during a lifetime exposure to a substance present as a contaminant in aquatic food sources 

(CCME, 1998). TRGs are calculated from the most sensitive of the 

available toxicity tests and applied to the Canadian wildlife species 

with the highest food intake/body weight ratio (CCME, 1998). TRGs 

are available for two organochlorine pesticides (DDT and toxaphene), 

which were derived using Wilson’s storm petrel and the mink as representative wildlife species (CCME, 1999a,b).

New York fish flesh criteria (FFC) for protection of piscivorous wildlife, noncancer values—These are intended to protect specific wildlife species from adverse effects other than cancer, such as mortality, 

reproductive impairment, and organ damage (Newell and others, 

1987). The New York State Department of Environmental Conservation 

(NYSDEC) derived these criteria using the same extensive laboratory animal toxicology database that is used to derive criteria for the 

protection of human health. Instead of extrapolating from laboratory 

animals to humans, the NYSDEC extrapolated from laboratory animals 

to wildlife. To represent birds and mammals, the NYSDEC selected a 

generic bird (with a body weight of 1 kg and a food consumption rate 

of 0.2 kg/day) and the mink. New York FFC are available for 8 of the 12 

organochlorine pesticides and pesticide groups measured by NAWQA 

in fish.

Proposed criteria from the Contaminant Hazard Review series—

Proposed tissue-based criteria for wildlife are included among recommendations for protection of natural resources in the Contaminant 

Hazard Review series developed by the U.S. Fish and Wildlife Service. 

Proposed criteria are available from this series for two organochlorine pesticides, toxaphene and chlordane. For chlordane (Eisler, 1990), 

the criterion is based on birds only—Eisler noted that criteria for protection of mammalian wildlife were lacking, and criteria for birds were 

incomplete and still required NOAELs from lifetime exposures. Wildlife 

benchmarks for toxaphene (Eisler and Jacknow, 1985) are based on 

criteria for human-health protection (ranging in various foods from 

0.1 to 7.0 mg/kg), which are expected to protect sensitive species of 

wildlife. 

Application of Fish-eating Wildlife Benchmarks 

for Fish

Fish-eating wildlife benchmarks for fish tissue were compared 

with concentrations of organochlorine pesticide compounds or 

groups measured by NAWQA in composite samples of whole fish 

(one sample per site). Concentrations measured by NAWQA were 

compared with both the lowest and the highest benchmark values 

available for each pesticide compound and group. The analysis thus 

reflects the degree of uncertainty in estimating the potential for 

adverse effects on wildlife. 

Wildlife Benchmarks for Pesticides in Whole Fish

Potential for Effects on Human Health, Aquatic Life, and Wildlife 109

Chapter 6
Most of the concentrations that exceeded 

a benchmark, as well as most of the variance 

between high and low benchmarks, were due to 

total DDT. Using the high and low ends of the 

range of benchmark values available for different 

pesticides (or pesticide groups), wildlife benchmarks were exceeded at 11 to 88 percent of urban 

stream sites for total DDT, 15 to 18 percent for 

dieldrin, and 0 to 10 percent for total chlordane. 

In agricultural streams, total DDT exceeded 

wildlife benchmarks at 29 to 87 percent of sites, 

dieldrin at 7 to 11 percent, and toxaphene at 0 to 

9 percent (fig. 6–18).

Percentage of stream sites

exceeding the benchmark

Percentage of stream sites

exceeding the benchmark

High benchmark

Low benchmark

Agricultural streams

Organochlorine compounds with concentrations 

greater than a wildlife benchmark

Urban streams

Total DDT

Toxaphene

Total chlordane

Endrin

Dieldrin

Total DDT

Dieldrin

Total chlordane

Total heptachlor

Toxaphene

6_15_bars

0 25 50 75 100 0 25 50 75 100

Figure 6–18. Contributions of individual organochlorine pesticide compounds or groups to exceedances 

of whole-fish tissue benchmarks for fish-eating wildlife show the potential significance of total DDT and 

dieldrin. Water-quality benchmarks are provided in Appendix 3B. 

Organochlorine compounds from historical pesticide use 

are still a concern for fish-eating wildlife in some streams 

(Photograph by W.H. Mullins © 1974).

The wildlife screening-level assessment for 

organochlorine compounds in fish tissue indicates that these compounds still occur at some 

sites at concentrations that have the potential to 

adversely affect fish-eating wildlife. Although 

there is relatively high uncertainty in benchmark 

values, total DDT and dieldrin accounted for 

most benchmark exceedances, and there were 

34 percent of agricultural sites and 25 percent 

of urban sites with concentrations that exceeded 

both low and high benchmark values for one or 

more pesticide compounds or groups. 

110 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 6

Emerging Issues for Assessment of 

Pesticide Effects

Although pesticides are among the most 

intensively studied of environmental contaminants, and many studies of fate and effects are 

required to register a pesticide for use, comprehensive assessment of their potential effects continues to present challenges. Two issues receiving 

particular attention by the scientific and regulatory communities are the potential effects of 

pesticide mixtures, the occurrence of which was 

examined in Chapter 5, and the potential effects 

of pesticides on endocrine systems.

Approaches for Assessing Potential 

Effects of Pesticide Mixtures on 

Humans and Aquatic Life

Understanding the potential effects of 

chemical mixtures on humans and the environment is one of the most complex problems facing 

scientists and regulatory agencies. USEPA identified this issue as a priority in its research strategy 

for 2000 and beyond (USEPA, 2000b). Although 

guidelines and detailed procedures for evaluating 

potential effects from exposure to chemical mixtures have been provided by USEPA (USEPA, 

1986, 2000b) and other agencies (ATSDR, 

2004b), implementation has been difficult 

because of the complexity of mixtures that occur 

in the environment and the inadequacy of data 

on the toxicity of the mixtures. Most toxicological testing is performed on single chemicals—

usually at high exposure levels—whereas most 

human and ecological exposures are to chemical 

mixtures at relatively low doses (USEPA, 2000b; 

ATSDR, 2004b).

Humans can be exposed to mixtures of 

pesticides and their degradates that occur in 

streams and ground water if such water is used as 

a source of drinking water and if treatment does 

not eliminate the pesticide compounds. Aquatic 

organisms are exposed to mixtures that occur 

in streams. Pesticide mixtures may be derived 

from common sources (such as point sources) or 

from multiple nonpoint sources, and may include 

several different types of pesticide compounds 

with different mechanisms of toxicity. Although 

a review of recent research on the effects of 

pesticide mixtures is beyond the scope of this 

report, the present approaches taken by USEPA 

and other agencies for regulating and assessing 

pesticide mixtures provide an indication of present knowledge and information gaps.

Evaluation and management of potential 

risks to humans of pesticide mixtures that may 

occur in drinking water are primarily addressed 

at the Federal level by USEPA and the Agency 

for Toxic Substances and Disease Registry 

(ATSDR). Much of the attention to potential 

effects of chemical mixtures on human health has 

been associated with risk assessments required 

for hazardous waste sites as part of implementing the Comprehensive Environmental Recovery, 

Compensation, and Liability Act (CERCLA), but 

specific assessment of pesticide mixtures is also 

now occurring to meet requirements of the Food 

Quality Protection Act (FQPA) of 1996. Under 

the FQPA, USEPA must assess the cumulative 

risks of pesticides that share a common mechanism of toxicity, or act the same way in the body. 

These cumulative assessments consider exposures from food, drinking water, and residential 

sources. USEPA also incorporates regional 

exposures from residential and drinking-water 

sources to account for the considerable variation in potential exposures across the country. 

To date, USEPA has determined that within each 

of four different chemical classes (organophosphates, N-methyl carbamates, triazines, and 

chloroacetanilides), several specific pesticide 

compounds have a common mechanism of toxicity and require cumulative risk assessments to 

better define the potential effects of exposure of 

humans to multiple pesticides within each class. 

The potential effects of chemical mixtures 

on aquatic life have not received as much attention as for human health, although USEPA’s 

Office of Research and Development, National 

Center for Environmental Assessment, has completed ecological risk-assessment guidelines that 

support the cumulative risk-assessment approach 

(USEPA, 2003f). The pesticide registration and 

reregistration processes require ecological risk 

assessment, which includes evaluation by USEPA 

of the likelihood that exposure to more than one 

pesticide and its degradates may cause harmful 

ecological effects.

Potential effects of pesticide mixtures on 

aquatic life also may be considered as part of 

assessments for National Pollutant Discharge 

Elimination System (NPDES) permits or hazardous waste sites. Procedures developed by USEPA 

for conducting assessments for NPDES permits 

involve a battery of tests, referred to as “whole 

effluent toxicity” (WET) tests, for both effluents 

and receiving waters. The WET tests are toxicPotential for Effects on Human Health, Aquatic Life, and Wildlife 111
ity tests applied to actual or simulated effluent 

and receiving water and, therefore, assess the 

combined toxicity to aquatic life of all contaminants present in water (USEPA, 2004f). Although 

the WET test procedures provide a methodology 

for directly testing ambient waters that contain 

mixtures, they have not yet been applied more 

broadly to assess mixtures of pesticides from 

nonpoint sources that do not involve NPDES 

permits. Similarly, the risk-assessment methods 

developed for mixtures that occur at hazardous 

waste sites (USEPA, 2003f) provide a systematic approach to assessing potential effects of 

pesticide mixtures on aquatic life, but they are 

generally not applied to ambient water-quality 

conditions.

In addition to these various approaches to 

addressing mixtures as part of the regulatory 

process, researchers are studying the effects of 

specific mixtures of pesticides and degradates 

and relating the occurrence of mixtures to their 

potential effects on aquatic ecosystems. The 

accompanying sidebar by Lydy and Belden 

(p. 114) provides a perspective on current understanding and the status of research regarding the 

potential effects of pesticide mixtures on aquatic 

life. NAWQA has begun to examine relations 

between biological measures of stream quality 

and the range of stresses introduced by agricultural and urban activities, including exposure 

to pesticides. The accompanying sidebar on the 

Pesticide Toxicity Index (p. 116) summarizes 

how the index is used by NAWQA as a relative 

indication of the potential toxicity of a mixture 

to aquatic life and illustrates its applications with 

examples from NAWQA studies. 

Although an array of approaches has been 

developed for assessing the potential effects of 

mixtures using the best available data on exposure and effects, progress toward understanding 

the potential effects of pesticide mixtures on 

humans and aquatic life has been hampered, 

in part, by sparse data on the composition and 

concentrations of mixtures that actually occur 

in streams and ground water. As examined in 

Chapter 5, pesticide degradates are potentially 

important components of pesticide mixtures that 

need to be considered when evaluating potential 

effects. Improved data on the occurrence and 

composition of mixtures from NAWQA and other 

studies can help to characterize the potential 

exposure of humans, aquatic life, and wildlife to 

mixtures and provide a basis for systematically 

prioritizing mixtures that may occur in streams 

and ground water. 

Endocrine Disruption and Pesticides

Endocrine systems are present in mammals, birds, fish, and other organisms. They are 

comprised of glands that produce hormones, 

which act as chemical messengers, and receptors 

in various organs and tissues that recognize and 

respond to the hormones. The endocrine system 

regulates all body functions from conception 

through adulthood, including the development 

of the brain and nervous system, the growth and 

function of the reproductive system, and metabolism and blood-sugar levels. Disruption of the 

endocrine system by a contaminant can occur in 

a number of ways, such as by mimicking a natural hormone, blocking the effects of a hormone, 

or causing overproduction or underproduction of 

hormones (Gross and others, 2003). 

More than 50 synthetic chemical compounds, including a number of pesticides, have 

been identified as potential endocrine disruptors 

in various studies over the past several years 

(National Academy of Sciences, 1999). The 

studies include bioassays demonstrating estrogenic or anti-estrogenic activity and field studies 

correlating contaminants with hormone-related 

effects. Examples of such field studies include 

feminization of gull embryos linked to elevated 

DDT (Fry and Toone, 1981), population declines 

of alligators in some Florida Lakes with elevated 

concentrations of organochlorine pesticides 

(including DDT) (Guillette and others, 1994), 

and feminization of fish in water bodies receiving municipal discharges or industrial effluents 

(Purdom and others, 1994).

In 1994, the NAWQA Program investigated 

the potential influence of contaminants on sex 

steroid hormones and other biomarkers in common carp (Goodbred and others, 1997). Abnormal ratios of sex steroid hormones in both male 

and female carp were found at some sites, and the 

ratio of estrogen to testosterone, an indicator of 

potential abnormalities in the endocrine system, 

was significantly lower at sites where some of the 

highest pesticide concentrations were detected 

(fig. 6–19). Further investigation is needed to 

determine whether (1) reduced hormone ratios 

are caused by pesticides, and (2) the reduced hormone ratios are associated with significant effects 

on fish populations.

The 1996 Food Quality Protection Act 

requires USEPA to screen and assess pesticides and other environmental contaminants for 

potential effects on human endocrine systems, an 

assessment which USEPA is extending to wild112 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 6

life as well. A review of NAWQA pesticide data 

compared with a list of potential endocrine disrupting compounds (Keith, 1997) indicates that 

17 pesticides measured by NAWQA in water are 

possible endocrine disruptors (USEPA has not 

yet designated pesticides that it considers to be 

potential endocrine disruptors). Eleven of these 

pesticides were among those most frequently 

found in NAWQA stream samples (fig. 6–20).

Research on the effects of chemicals on 

endocrine systems is in its relatively early stages. 

Several important aspects are still unclear, 

including the degree to which such effects occur 

in the environment; whether effects on individual 

organisms translate to effects on populations and 

communities; and at what concentrations effects 

on populations become significant. There is considerable scientific uncertainty about the causes 

of reported effects (Kavlock and others, 1996). 

A major effort is underway by USEPA and other 

agencies to systematically identify and better 

understand endocrine disruptors (USEPA, 1998).

5

4.5

4

3.5

3

2.5

2

1.5

1

0.5

0

0 0.5 1 1.5 2 2.5 3 3.5

Estrogen/Testosterone ratio

Total concentration of pesticides in water,

in micrograms per liter

Correlation of carp hormone levels

with pesticides in water

Females

Males

Agricultural streams

Potential endocrine disruptors in water

Urban streams

Percentage of time detected

Atrazine

Metolachlor

Alachlor

Metribuzin

Trifluralin

Simazine

2,4-D

Chlorpyrifos

Carbaryl

Malathion

Dieldrin

0 25 50 75 100 0 25 50 75 100

Insecticides Herbicides

Figure 6–19. Ratios of estrogen-to-testosterone 

in carp from 11 streams sampled by NAWQA in 

1994 were inversely correlated with pesticide 

concentrations. Low ratios indicate potential 

abnormalities in carp endocrine systems (Goodbred 

and others, 1997). 

Figure 6–20. Eleven pesticides that have been identified as 

potential endocrine disruptors (Keith, 1997) were among the 

pesticides most frequently detected in NAWQA water samples 

from agricultural and urban streams.

Potential for Effects on Human Health, Aquatic Life, and Wildlife 113
NAWQA studies show that the most common form of pesticide exposure for aquatic organisms is simultaneous exposure to 

multiple pesticides. More than 50 percent of all stream samples 

contained five or more pesticides. Yet, most pesticide research, 

historically and currently, has evaluated the effects of individual 

pesticides as if they occurred alone. Scientists and risk assessors 

are only in the beginning stages of developing the knowledge base 

and procedures for evaluating the potential environmental effects 

of pesticide mixtures in aquatic ecosystems.

Conceptual Models of Mixture Effects

Research on mixtures indicates that a wide array of possible 

interactions among pesticides may occur, but they all fall into one 

of four categories:

Independent—Co-occurring pesticides act independently of one 

another, with each causing the degree of effects on a population 

as would be expected from its individual concentration. This might 

occur for pesticides with different target organs and modes of 

action. 

Additive—Co-occurring pesticides act in an additive manner, with 

effects on a population as would be expected by summing the 

toxicity-normalized concentrations of multiple individual pesticides 

that are present. This might be expected for pesticides with similar 

chemical structures and a common mode of action.

Antagonistic—Co-occurring pesticides have a combined toxicity 

that is less than that predicted from the additive model. 

Synergistic—Co-occurring pesticides have a combined toxicity 

that is greater than that predicted from the additive model.

The additive model, also called the Concentration Addition 

Model, is the most common baseline used for assessing effects 

of pesticide mixtures, although not all mixtures strictly follow it. In 

a 2-compound mixture, the concentration of chemical A and the 

concentration of chemical B would be normalized (weighted) by 

toxicity as follows: the concentration of each chemical present in 

the sample is divided by its toxicity value (usually the concentration needed to cause a 50-percent effect in a population) and the 

toxicity-weighted concentrations are then added together. The 

effect expected would then be based on this normalized total 

concentration. For example, if two pesticides that have the same 

toxicity are each present in a stream at 10 μg/L, then the expected 

effect would be the same as the effect of 20 μg/L of either one of 

the compounds alone.

Experimentally, additive toxicity has been observed for several groups of mixtures, including 2-compound mixtures of the 

s-triazine herbicides atrazine and cyanazine in reproductive tests 

with the green alga Chlorella fusca (Faust and others, 1993) and 

2-compound mixtures of several organophosphate insecticides, 

including chlorpyrifos, diazinon, and azinphos methyl, in tests with 

midges (Lydy and Austin, 2004). In addition, the organophosphate 

insecticides chlorpyrifos and diazinon were strictly additive in their 

toxicity to the cladoceran, Ceriodaphnia dubia, in toxicity studies 

performed in natural, storm, and laboratory waters (Bailey and 

others, 1996, 1997, 2000). 

Several studies have shown that pesticide interaction can 

result in less (antagonistic) or more (synergistic) toxicity than 

predicted by the Concentration Addition Model. For example, 

researchers have demonstrated that simultaneous exposure to 

esfenvalerate (a pyrethroid insecticide) and diazinon (an organophosphate insecticide) resulted in greater than additive toxicity 

to fathead minnows (Denton and others, 2003). The likely reason 

for this synergism is that diazinon inhibits the esterase enzymes, 

thus reducing the organism’s capability to detoxify pyrethroids. 

Other studies have shown that the herbicide atrazine, when present at concentrations above 40 μg/L, increases the toxicity of the 

organophosphate insecticides chlorpyrifos (fig. 6–21) and diazinon 

to aquatic invertebrates (Belden and Lydy, 2000; Anderson and 

Lydy, 2002). Note that atrazine itself is not acutely toxic to these 

invertebrates, even at high concentrations in water. In this case, 

the reason for the increased toxicity is that atrazine induces 

(increases production of) specific oxidative enzymes, resulting in a 

higher transformation rate of chlorpyrifos into a more toxic metabolic product (Belden and Lydy, 2000). In both of these examples, 

one contaminant changed the organism’s capacity to metabolize 

the other contaminant, thus increasing or decreasing the amount 

of pesticide or pesticide breakdown products within the organism, 

and leading to large changes in the degree of toxicity.

100

80

60

40

20

0

Chlorpyrifos

0.17 ppb

Atrazine

200 ppb

Chlorpyrifos

0.17 ppb

and atrazine

40 ppb

Chlorpyrifos

0.17 ppb

and atrazine

80 ppb

Chlorpyrifos

0.17 ppb

and atrazine

200 ppb

Immobility, in percent

Toxicity of mixtures of chlorpyrifos and atrazine

Figure 6–21. Although atrazine itself was not acutely 

toxic to the aquatic invertebrates tested in this study 

(chironomids), an increase in atrazine concentration 

caused an increase in the toxicity observed for chlorpyrifos 

(a synergistic interaction), as indicated by increased 

immobility (Belden and Lydy, 2000). (Concentrations are 

shown in ppb [parts per billion] as in the original report, 

which is equivalent to micrograms per liter.)

114 Pesticides in the Nation’s Streams and Ground Water, 1992–2001

Assessing Potential Effects of Pesticide Mixtures 

Michael J. Lydy and Jason B. Belden

Fisheries and Illinois Aquaculture Center and Department of Zoology

Southern Illinois University
Table 6–2. Selected studies of pesticide mixtures containing diazinon illustrate the spectrum of possible responses for such mixtures. 

The types of compounds included are two organophosphate insecticides (OP), a pyrethroid insecticide (P), a triazine herbicide (T), and 

a nutrient. 

Mixture Type of 

compound Species tested Result

Deviation from 

concentration 

addition

Reference

Diazinon and chlorpyrifos OP

OP

Midges

(aquatic invertebrate)

Additive None Lydy and Austin (2004)

Diazinon and chlorpyrifos OP

OP

Ceriodaphnia dubia

(aquatic invertebrate)

Additive None Bailey and others (1996, 1997)

Diazinon and esfenvalerate OP

P

Fathead minnows Synergistic 140 to 170 percent 

greater toxicity

Denton and others (2003)

Diazinon and atrazine OP

T

Midges, amphipods 

(aquatic invertebrates)

Synergistic Up to 400 percent 

greater toxicity

Anderson and Lydy (2002); 

Belden and Lydy (2000)

Diazinon and ammonia OP

Nutrient

Ceriodaphnia dubia

(aquatic invertebrate)

Antagonistic 27 to 32 percent 

less toxic

Bailey and others (2001)

Studies of the toxicity of pesticide mixtures have resulted in the 

full spectrum of additive, synergistic, and antagonistic responses. 

Generally, pesticides within the same pesticide class and that have 

similar structures and a common mode of action (for example, 

organophosphate insecticides) are more likely to follow the additive 

model, while pesticides from different classes (for example, herbicides and insecticides) have more varied effects. Table 6–2 summarizes the results from selected studies of mixtures containing the 

organophosphate insecticide diazinon. Because of the complexity of 

the modes of action and chemical transformations that occur for each 

pesticide, the toxicity of most pesticide mixtures will deviate from the 

simple additive model. It is not known how likely such deviations from 

additivity are, nor is there consensus on how large a deviation from 

the model is significant. In many cases, this deviation may be smaller 

than that obtained from testing the organisms under slightly different 

conditions (intraspecies toxicity testing), indicating that other sources 

of uncertainty may be more significant than errors in mixture models. 

However, until a more thorough understanding of pesticide interactions is achieved, the possibility of pesticide combinations resulting 

in greater toxicity than predicted by the additive model needs to be 

considered. 

Implications

In most situations, a mixture of pesticides presents a greater risk 

to aquatic organisms than do any of the individual components of 

the mixture. The ecological effects caused by mixtures of pesticides, 

however, are highly uncertain and are in the relatively early stages of 

investigation. Further research must be conducted before the possible impacts that pesticide mixtures may have on the environment can 

be determined. The large numbers of chemicals and varying exposure 

routes that occur in the environment make testing every possible 

exposure scenario impossible. For example, in a mixture of 20 compounds, there are 190 pairs of compounds, and more than a million 

possible combinations (pairs, triples, and so on). Thus, it makes sense 

for researchers assessing mixture effects to prioritize and test those 

combinations with a high probability of environmental occurrence and 

those that are useful in developing refined models to predict the toxicity of similar pesticide mixtures.

Ultimately, aquatic toxicologists need to understand the dynamic 

world that organisms encounter. Besides pesticides, organisms are 

exposed to other types of chemical contaminants (such as metals and 

industrial contaminants) and also biological and physical stressors 

(such as changes in flow rate, temperature, habitat, food availability, and predation) simultaneously. It is likely that these stressors 

interact. However, until we better understand the biology of aquatic 

systems, from the molecular to the ecosystem level, we will continue 

to struggle in predicting the existence and significance of chemical 

interactions.

Potential for Effects on Human Health, Aquatic Life, and Wildlife 115

Chapter 6
Rank of tolerant invertebrate species

Decreasing quality

Rank correlation = 0.74

Pesticide Toxicity Index applied 

to agricultural streams

0 5 10 15 20 25

0

5

10

15

20

25

Rank of PTI

Increasing PTI

0

0

5

10

15

20

25

30

5 10 15 20 25 30

Rank of sensitive invertebrate species

Decreasing quality

Rank of PTI

Increasing PTI

Rank correlation = –0.60

Pesticide Toxicity Index applied 

to urban streams

116 Pesticides in the Nation’s Streams and Ground Water, 1992–2001

Pesticide Toxicity Index

To expand the assessment of potential effects of pesticides in 

stream water on aquatic life, NAWQA developed a Pesticide Toxicity Index (Munn and Gilliom, 2001). The Pesticide Toxicity Index (PTI) 

accounts for multiple pesticides in a sample, including pesticides 

without established benchmarks for aquatic life. The PTI combines 

information on exposure of aquatic biota to pesticides (measured 

concentrations of pesticides in stream water) with toxicity estimates 

(results from laboratory toxicity studies) to produce a relative index 

value for a sample or stream. The PTI value is computed for each 

sample of stream water by summing the toxicity quotients for all pesticides detected in the sample. The toxicity quotient is the measured 

concentration of a pesticide divided by its toxicity concentration from 

bioassays (such as an LC50 or EC50). For each sample, separate PTI values are computed for fish and benthic invertebrates. This approach 

follows the Concentration Addition Model of toxicity described by 

Lydy and Belden (accompanying sidebar, p 114). Although simple 

additivity is unlikely to strictly apply for complex mixtures of pesticides 

from different classes and with different effects and modes of action, 

the PTI is still useful as a relative index. Deneer (2000) reported that 

“for more than 90 percent of 202 mixtures in 26 studies, concentration 

addition was found to predict effect concentrations correctly within 

a factor of two.” While the PTI does not indicate whether water in a 

sample is toxic, its value can be used to rank or compare the relative 

potential toxicity of different samples or different streams.

The PTI provides a means to rank different stream sites compared with each other and is a tool for investigating relationships 

between pesticide levels and the quality of aquatic ecosystems. For 

example, pesticides were commonly detected in agricultural streams 

and drains throughout the Yakima River Basin, often at concentrations exceeding one or more aquatic-life benchmarks for individual 

pesticides (Fuhrer and others, 2004). Data for 24 stream sites in the 

Yakima River Basin showed that the number of pollution-tolerant 

Figure 6–22. Streams and drains in the Yakima River 

Basin with the highest PTI values tended to have 

the highest numbers of pollution-tolerant benthic 

invertebrates, indicating lower water quality. The 

ranks were significantly correlated at a 95-percent 

confidence level. (Modified from Fuhrer and others, 

2004.)

Figure 6–23. Streams in the Dayton and Cincinnati, 

Ohio areas with the highest PTI values tended to 

have the lowest numbers of sensitive invertebrate 

species, indicating lower water quality. The ranks 

were significantly inversely correlated at a 95-

percent confidence level. (Modified from Rowe and 

others, 2004.) 

116 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
0

10-6 10-5 0.0001 0.001

Pesticide Toxicity Index

0.01 0.1 1

10-6 10-5 0.0001 0.001

Pesticide Toxicity Index

0.01 0.1 1

20

40

60

80

100

0

20

40

60

80

100

Agricultural

Urban

Undeveloped

Mixed

Land use

Frequency that stream samples equaled or exceeded the indicated PTI value

6_19

Invertebrates

Pesticide Toxicity Index for NAWQA stream sites

Fish

Potential for Effects on Human Health, Aquatic Life, and Wildlife 117

Figure 6–24. Invertebrate PTI values for all land-use 

categories were more than 10 times higher than fish 

PTI values. Urban and agricultural streams had the 

highest PTI values. This analysis is based on the 95th 

percentile PTI value for each site, which is an estimate 

of the PTI value that was exceeded by 5 percent of 

samples at the site.

benthic invertebrates (higher numbers indicate a stressed ecosystem) 

significantly increased with the rank of PTI values (fig. 6–22). Pesticides, however, are only one of many factors that may affect aquatic 

communities—other factors include physical habitat quality, food 

availability, and the presence of other contaminants. Detailed studies 

are required to distinguish the relative roles of different chemical and 

physical factors.

Another example is for streams in the Dayton and Cincinnati, Ohio, 

urban areas, which were studied in the Great and Little Miami River 

Basins (Rowe and others, 2004). Results for 30 streams with varying degrees of urban land use in their watersheds indicated that the 

number of sensitive invertebrates (lower numbers indicate a stressed 

ecosystem) significantly decreased with increasing PTI values 

(fig. 6–23). As with the Yakima River Basin example, this correlation 

does not demonstrate a cause-and-effect relationship between pesticides and the benthic invertebrate community. The PTI was one of 

several factors found to correlate with degree of urbanization—which 

also included chloride levels in water and synthetic chemicals in bed 

sediment—that may affect benthic invertebrates (Rowe and others, 

2004).

For a national-level perspective, the PTI was used to rank NAWQA 

stream sites by the potential toxicity of measured pesticide concentrations to fish and benthic invertebrates. Invertebrate PTI values 

generally were more than 10 times higher than those for fish, as 

shown by frequency distributions of the 95th percentile PTI values for 

streams in all land-use settings (fig. 6–24). The higher toxicity values 

for invertebrates reflect greater sensitivity of invertebrates compared 

with fish, particularly to insecticides. A large proportion of benthic 

invertebrates are insects, which explains the high relative toxicity 

of insecticides to this taxonomic group. PTI values for both fish and 

invertebrates are highest for samples collected from agricultural and 

urban streams, lowest for undeveloped streams, and intermediate for 

mixed-land-use streams. These results are consistent with the results 

of the screening-level assessment (using aquatic-life benchmarks) of 

the potential effects of pesticides in water on aquatic life.

Chapter 6

Potential for Effects on Human Health, Aquatic Life, and Wildlife 117

Chapter 7

Strategies for managing pesticides require far more informa-

tion than we can afford to directly measure for all the places, 

times, and pesticides of interest. In addition, many strategic 

decisions—such as setting monitoring priorities, approving 

a pesticide registration, and determining how much to spend 

on a management strategy—inherently depend on predicting 

the potential effects of pesticides on water quality for loca-

tions that have never been directly assessed. In these situations, 

statistical models and other types of models can be useful for 

predicting water-quality conditions at unmonitored locations 

under a range of possible circumstances. Such tools are essen-

tial for efficient water-quality management.

In this chapter, three examples 

illustrate the development 

of statistical models from 

NAWQA data and some of 

the ways in which the models 

can be applied to nationalscale analysis of water-quality 

conditions.

7

Prediction Where Data are Inadequate Photograph by Albert E. Becher
Approach to Prediction

NAWQA pesticide data collected from 1992 

to 2001 support the development and testing of 

a wide range of models, particularly statistical 

models. Statistical models have been developed 

by NAWQA to predict pesticide levels in streams 

and ground water for locations where pesticide 

concentrations have not been measured. This 

expansion of water-quality assessment from 

individual monitoring sites to unstudied locations by use of models for prediction, or spatial 

extrapolation, is fundamental to extending the 

targeted local and regional studies of NAWQA to 

a comprehensive national assessment (Alaska and 

Hawaii have not been included in these models 

because there are no suitable pesticide-use data 

for these States).

The NAWQA statistical models for pesticides use linear regression methods to establish 

quantitative linkages between pesticide concentrations measured at NAWQA sampling sites 

and a variety of anthropogenic (human-related) 

and natural factors that affect pesticides. Such 

factors include pesticide use, soil characteristics, 

hydrology, and climate—collectively referred 

to as explanatory variables. Model-development 

data consist of measured pesticide concentrations or detection frequencies, together with the 

associated values of the explanatory variables for 

the sampling sites. The models are built using 

the explanatory variables that best correlate with, 

or explain, the concentrations or frequencies of 

occurrence of pesticides observed in streams and 

ground water. Although explanatory variables 

included in the models are significantly correlated with pesticide concentrations or detection 

frequencies, the specific cause-and-effect relations responsible for the observed correlations are 

not always clear, and inferences regarding causes 

should be considered as hypotheses. 

In developing the pesticide models, all 

potential explanatory variables were required to 

have values available from existing data sources 

for all locations in the conterminous United 

States, so that national extrapolation would be 

possible (the only exception, as explained below, 

was fish lipid content for the dieldrin model). 

Overall, 30 to 60 possible variables were considered, depending on the specific model; these 

were reduced to the 4 to 6 explanatory variables 

that were most significant and yielded optimal 

model formulations. Each model incorporates an 

uncertainty analysis, which allows assessment of 

the reliability of the model predictions and also 

the expression of model predictions as probabilities that concentrations will exceed a specific 

value, such as a water-quality benchmark, at a 

particular location.

The three NAWQA models and nationally 

extrapolated results presented in this chapter are 

those developed for (1) concentrations of atrazine 

in stream water; (2) concentrations of dieldrin 

in whole fish; and (3) detection frequencies for 

atrazine in shallow ground water underlying agricultural settings. The extrapolations for atrazine 

concentrations in stream water and dieldrin concentrations in fish tissue are for streams included 

in the USEPA River Reach file (Nolan and 

others, 2003), which includes more than 600,000 

miles of streams and more than 60,000 individual 

stream reaches with watersheds. The extrapolations of detection frequencies for atrazine in 

shal­low ground water were made for all areas of 

the Nation where at least 50 percent of the land 

is in agricultu­ral use. More detailed information 

on model development methods and supporting 

data, as well as uncertainty analyses, are provided by Larson and others (2004), Nowell and 

others (2006), and Stackelberg and others (2006). 

Addi­tional work is currently underway on (1) a 

multi-pesticide model for stream water that incorporates selected chemi­cal and physi­cal properties 

of each compound, (2) expanding the models for 

fish tissue to include additional organo­chlorine 

compounds, and (3) site-specific, concentrationbased models for atrazine in ground water. 

Atrazine Concentrations in Streams 

Model predictions of atrazine levels in 

streams across the Nation show the highest 

annual mean concentrations throughout the highuse areas of the Corn Belt and the Mississippi 

Valley and Delta regions, and in some areas of 

Texas, Pennsylvania, and Maryland. Figure 7–1 

shows measured concentrations used to develop 

the model and figure 7–2 shows predicted concentrations. As noted along with other model 

details in the accompanying sidebar (p. 121), 

the model is based on the time-weighted annual 

mean for each model-development site. Annual 

means for a few streams in the Ohio and Mississippi Valleys and in southern Louisiana are 

predicted to exceed 3 μg/L, the human-health 

benchmark used for atrazine (Chapter 6 and 

Appendix 3A). The benchmark for atrazine is the 

USEPA MCL for drinking water. As a drinking-water standard, the MCL applies to finished 

water in public water supplies, whereas the 

predictions shown in figure 7–2 are for untreated 

120 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 7

Development and Application of the Atrazine 

Model for Stream Water

As described by Larson and others (2004), the model for estimating atrazine concentrations in streams is based on time-weighted annual mean 

concentrations measured by NAWQA from 1992 to 2001 at 112 sites (fig. 7–1). 

The single most complete year of data was used to calculate the annual 

mean concentration for each site. The predicted values in figure 7–2 are 

median estimates of the annual mean, such that 50 percent of the actual 

annual means are expected to be greater than, and 50 percent less than, 

the predicted value. Nonagricultural uses of atrazine are not included and, 

as a result, predictions may represent underestimates for watersheds with 

substantial nonagricultural use. To illustrate a practical example of how such 

models can be applied to water-quality assessment, model estimates are 

compared with the human-health benchmark for atrazine, USEPA’s Maximum 

Contaminant Level (MCL) for drinking water (Chapter 6 and Appendix 3). The 

model also was used to estimate the probability, after accounting for model 

uncertainty, that any particular stream site may have an annual mean atrazine 

concentration greater than 3 μg/L (fig.7–2 ). 

 0.03

0.03 – 0.3

 0.3 – 3

 3

Annual mean atrazine concentration,

in micrograms per liter

Model development sites

Figure 7–1. The model for annual mean concentrations of atrazine 

in streams was developed from data for 112 sites distributed across 

the country, which represent a wide range of hydrologic settings and 

atrazine concentrations.

stream water. Comparisons of model predictions 

with human-health benchmarks, however, serve 

as screening-level assessments of the suitability 

of potential drinking-water sources, as discussed 

in Chapter 6.

For more than half of the streams with 

a predicted annual mean atrazine concentration exceeding 0.3 μg/L (fig. 7–2), there is at 

least a 5-percent chance that the actual annual 

mean concentrations will exceed the humanhealth benchmark of 3 μg/L (fig. 7–2). Model 

estimates of probabilities shown in figure 7–2 

indicate that at least 1 out of 20 (5 percent) of the 

streams shown in yellow, orange, or red would be 

expected to have annual mean atrazine concentrations greater than 3 μg/L. These streams may not 

be suitable as sources of drinking water without 

treatment or other management strategies to 

reduce atrazine concentrations. The streams with 

a greater than 5-percent probability of exceeding 

the benchmark represent about 7 percent of the 

Nation’s stream miles (45,704 of 649,935 mi). 

Approximately 192 stream miles (less than 

1/10th of 1 percent of the Nation’s stream miles) 

are predicted to have more than a 50-percent 

probability of exceeding 3 μg/L. 

The model indicates that atrazine use 

intensity is the most important factor explaining 

atrazine concentrations in streams—the more 

intensive the use of atrazine in a watershed, the 

higher the atrazine concentration in the stream. 

Specifically, estimated atrazine use intensity 

within each watershed explains 64 percent of 

the variance in annual mean atrazine concentrations in streams across the Nation. Four additional variables explain another 13 percent of 

the variability, most of which is accounted for 

by rainfall erosivity and soil erodibility—factors 

used in the revised Universal Soil Loss Equation 

(Renard and others, 1997). Rainfall erosivity 

and soil erodibility quantify, respectively, the 

energy of storms in a specific area (averaged over 

several years), and the susceptibility of soils to 

erosion by runoff. As these two factors increase, 

atrazine concentrations also increase, indicating 

that transport of atrazine is highest in areas of 

high-energy rain storms and in areas where soils 

are most susceptible to erosion. Alternatively, 

soil erodibility may indicate high surface runoff, 

rather than actual transport of atrazine with soil 

particles. Overall, the complete model explains 

a total of 77 percent of the variance in observed 

annual mean atrazine concentrations. 

Prediction Where Data are Inadequate 121
Figure 7–2. Model predictions of annual mean atrazine concentrations in streams across the 

Nation show the highest concentrations (orange and red streams) throughout the high-use areas of 

the Corn Belt and the Mississippi Valley, and in some areas of Texas, Pennsylvania, and Maryland. 

Model predictions of the probability that atrazine concentrations are greater than the humanhealth benchmark of 3 μg/L for drinking water indicate that many streams in the Corn Belt and 

Mississippi Valley and Delta regions have greater than a 5-percent probability of having annual mean 

concentrations greater than the benchmark.

Predicted annual mean concentration of atrazine in streams

 0.03

0.03 – 0.3

0.3 – 3

 3

Predicted annual mean concentration of atrazine, in micrograms per liter

7_02ab_atrazine

Probability of exceeding the human-health benchmark for atrazine

 5%

5 – 25%

25 – 50%

 50%

Probability of exceeding 3 micrograms per liter

122 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 7

Development and Application of the Dieldrin Model 

for Whole Fish

As described by Nowell and others (2006), the model for estimating dieldrin 

concentrations in fish is based on concentrations measured in whole fish sampled by NAWQA from 1992 to 2001 at 648 sites across the country. The 514 sites 

shown in figure 7–3 are limited to the subset of model development sites with 

whole-fish samples having 2.3–10.4 percent lipid content (the lowest and highest 

10 percent of lipid levels were excluded from the map, but not model development). One composite sample (each composed of 5–10 fish of a single species) 

was collected at each site. The national data include 59 different species of 

fish, most frequently common carp (29 percent of samples) and white sucker 

(26 percent). One effect of compositing is to reduce variability in contaminant 

concentrations caused by differences in age and size among individual fish. An 

important explanatory variable in the dieldrin model is fish lipid content, which is 

not nationally available for all streams because it is a characteristic of the fish, 

rather than the stream or watershed. The inclusion of fish lipid content in the 

model accounts—to some extent—for differences among fish in age, size, and 

species because lipid content generally varies among species and increases 

(within a species) with increasing fish age and size (Nowell and others, 1999).

Model predictions were made using the national average lipid content for 

whole fish, which was 6.2 percent for samples collected by NAWQA and the 

U.S. Fish and Wildlife Service (Schmitt and Bunck, 1995). Predicted concentrations of dieldrin in fish, shown in figure 7–4, are median estimates for fish with 

6.2 percent lipid content. Consequently, actual concentrations are expected to 

be lower than the predicted value at 50 percent of sites and higher at 50 percent 

of sites. Also, fish with lipid content greater than 6.2 percent would likely have 

higher dieldrin concentrations, and fish with lower lipid content would likely have 

lower dieldrin concentrations, than those shown in figure 7–4. As examples, lipid 

content values typical of common fish species in the United States are lake trout, 

15 percent; channel catfish, 7.5 percent; common carp, 6.5 percent; white sucker, 

5.8 percent; largemouth bass, 4.2 percent; and bluegill, 3.1 percent. See Nowell 

and others (2006) for further discussion of uncertainty in model predictions.

To illustrate a practical example of how such models can be applied to waterquality assessment, model estimates are compared with the New York guideline 

for the protection of fish-eating wildlife, which was the highest wildlife benchmark compiled for dieldrin in fish tissue (Chapter 6 and Appendix 3B). The model 

also was used to estimate the probability, after accounting for model uncertainty, 

that any particular stream site may have a dieldrin concentration greater than 

120 μg/kg in whole fish with a 6.2 percent lipid content (fig. 7–4).

Dieldrin Concentrations in Fish

Model predictions of dieldrin concentrations 

in whole fish show the highest concentrations 

in the Corn Belt—especially in Illinois—where 

aldrin was heavily used on cropland. Figure 7–3 

shows measured concentrations used to develop 

the model and figure 7–4 shows predicted 

concentrations. Dieldrin is an organochlorine 

compound that was used as an insecticide until 

its agricultural use was discontinued in the early 

1970s, and it is also a degradate of aldrin, another 

insecticide that was used for agricultural purposes through the early 1970s. As noted along 

with other model details in the accompanying 

sidebar, model predictions are for fish with a 6.2 

percent lipid content, the national average lipid 

content for all whole fish sampled. Most streams 

that are predicted to have a dieldrin concentration greater than 25 μg/kg (micrograms per 

kilogram of fish tissue, wet weight) also have 

a 5 percent or greater chance (more than 1 in 

20) of exceeding 120 μg/kg (fig. 7–4), which is 

a wildlife benchmark for dieldrin in fish tissue 

(120 μg/kg is the highest of the dieldrin benchmarks compiled for this report; see Chapter 6 and 

Appendix 3B). These streams represent about 

6 percent of the Nation’s stream miles (40,222 

out of 649,935 mi). Approximately 627 stream 

miles (about 1/10th of 1 percent of the Nation’s 

stream miles) are predicted to have a 50-percent 

or greater probability of exceeding the dieldrin 

wildlife benchmark of 120 μg/kg. 

The dieldrin model indicates that the amount 

of forested land in a watershed is the most important factor explaining the concen­trations of dieldrin observed in fish—the greater the proportion 

of forested land (where historical use would have 

been least), the lower the dieldrin in fish tissue. 

Fish lipid content was also an important variable, 

which is consistent with the fact that organochlorine pesticides are hydrophobic compounds, 

which have a strong affinity for lipids, and thus 

tend to accumulate in high-lipid tissues. Two 

additional factors in the dieldrin model that, like 

forested land, are related to past use of dieldrin 

and aldrin represent (1) the estimated historical 

use of the compounds in agriculture and (2) their 

use for termite control. Dieldrin concentra­tions 

decrease with in­creasing amounts of forested 

land and increase with increasing historical use in 

agriculture or for termite control. Together, these 

three use-related factors and lipid content explain 

58 percent of the variability in dieldrin concentrations measured in whole fish in streams across 

the Nation. With the addition of two other less 

influential variables, the complete model explains 

64 percent of this variability.

Figure 7–3. The 

model for dieldrin 

concentrations in 

whole-fish tissue 

was developed 

using data from 648 

sites throughout the 

Nation, representing a 

wide range of dieldrin 

concentrations (514 

sites with 2.3–10.4 

percent fish lipid are 

shown). Predictions 

were made for fish 

with a lipid content 

of 6.2 percent, the 

national average for 

whole fish. 

Dieldrin concentration, 

in micrograms per kilogram (wet weight)

 5

5 – 25

25 – 120

 120

Model development sites

Prediction Where Data are Inadequate 123
Figure 7–4. Model predictions of dieldrin concentrations in whole fish in streams across 

the Nation show the highest concentrations in the Corn Belt, particularly Illinois, where 

aldrin (which degrades to dieldrin) was heavily used on cropland. Model predictions of the 

probability that dieldrin concentrations exceed the wildlife benchmark of 120 μg/kg indicate 

that there is greater than a 5-percent probability in many Corn-Belt streams that whole fish 

(with 6.2 percent lipid) contain dieldrin concentrations that exceed the benchmark. 

Predicted dieldrin concentrations

in whole fish with 6.2 percent lipid content

Predicted dieldrin concentration, in micrograms per kilogram (wet weight)

 5

5 – 25

25 – 120

 120

Probability of exceeding 120 micrograms per kilogram

Probability of exceeding the wildlife benchmark for dieldrin

in whole fish with 6.2 percent lipid content

5%

5 – 25%

25 – 50%

50%

124 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 7

Atrazine Detection Frequencies in 

Shallow Ground Water

Model predictions show that the highest frequencies of atrazine detection in shallow ground 

water beneath agricultural areas are expected 

in parts of the western Corn Belt, eastern Great 

Plains, Pacific Northwest (eastern Washington), 

and Mid-Atlantic regions (especially southeastern Pennsylvania). Figure 7–5 shows measured detection frequencies used to develop the 

model and figure 7–6 shows predicted detection 

frequencies for each square kilometer of land 

with 50 percent or more agricultural land. The 

areas with the highest frequencies of detection 

are those with relatively high atrazine use in 

hydrologic settings that also favor the transport of 

pesticides to ground water.

In contrast to the model for atrazine concentrations in stream water, atrazine use is not 

the most important factor for predicting atrazine 

occurrence in ground water. This finding is consistent with results from an earlier study of relations between atrazine in ground water and various land-use factors by Kolpin (1997), in which 

atrazine use was not found to be significantly 

correlated with atrazine occurrence in ground 

water. In the model presented herein, atrazine use 

explains only about 7 percent of the overall variability in the frequency of its detection in ground 

water. The two most important factors were 

found to be the proportion of land with subsurface tile drain systems and other artificial drainage, and the average vertical permeability of soil, 

which together explain 48 percent of the variability in atrazine detection frequencies. As the 

amount of artificial drainage increases, predicted 

detection frequencies decrease—a finding consistent with the fact that artificial drainage systems divert water and pesticides away from the 

ground-water system. Conversely, as the average 

vertical permeability of soils increases, predicted 

detection frequencies also increase because water 

and pesticides at the land surface are more likely 

to move vertically to ground water in areas with 

high-permeability soils. The influential role of 

these factors is particularly evident in Indiana and 

Ohio, where atrazine use is intense, but NAWQA 

studies, like several other previous studies 

(Barbash and Resek, 1996), found relatively low 

atrazine detection frequencies in ground water. 

Soils in these areas tend to be poorly drained and 

require artificial drainage to dewater the agricultural fields, thus reducing atrazine transport to 

ground water. With the addition of two other less 

influential variables, the complete model explains 

58 percent of the variability in atrazine detection 

frequencies observed in shallow ground water 

beneath the agricultural areas studied.

Development and Application of the Atrazine 

Model for Ground Water

As described by Stackelberg and others (2006), the model for predicting 

atrazine occurrence in shallow ground water within agricultural areas is 

based on the frequencies of detection measured by NAWQA from 1992 to 2001 

in 52 studies, each of which sampled about 20 to 30 shallow wells in agricultural areas (fig. 7–5). The model was used to predict the frequency of atrazine 

occurrence in shallow ground water in agricultural areas of the United States 

(fig. 7–6). Predictions were made for each 1 square kilometer area with 50 

percent or more agricultural land use. Nonagricultural use of atrazine was not 

included in use estimates, and thus, predictions may underestimate occurrence in areas where nonagricultural use is substantial.

Figure 7–5. The model for atrazine occurrence in shallow ground water 

within agricultural areas was developed from frequencies of detection 

in wells sampled for studies of shallow ground water in 52 agricultural 

areas across the country. The ground-water studies represent a wide 

range of agricultural and hydrologic settings, as well as atrazine 

detection frequencies.

Detection frequency, as a percentage of wells 

sampled in each study

Model development sites

 25 

25 – 50

50 – 75

 75

Prediction Where Data are Inadequate 125
Figure 7–6. Model predictions 

show that the highest frequencies of 

atrazine detection in shallow ground 

water beneath agricultural areas are 

expected in parts of the western Corn 

Belt, eastern Great Plains, Pacific 

Northwest (eastern Washington), 

and Mid-Atlantic regions (especially 

southeastern Pennsylvania). These 

areas represent relatively high 

atrazine use in hydrologic settings 

that favor the transport of pesticides 

to ground water.

Predicted frequency of detections, as a percentage of shallow wells

25

25 – 50

50 – 75

75

7_06_gw_atrazine

Predicted occurrence of atrazine in ground water

No prediction — areas have less than

50 percent agricultural land use

126 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 8

Long-term trends and changes in pesticide concentrations 

in streams and ground water are controlled largely by shifts 

in pesticide use, but the rates and geographic distributions of 

changes are also influenced by the chemical and physical prop-

erties of the pesticides and characteristics of the hydrologic 

settings. The first decade of NAWQA assessments focused 

primarily on establishing baseline conditions for comparison 

with future measurements as part of a long-term approach to 

tracking trends. Many trends, particularly those on a national 

scale, cannot yet be evaluated because not enough time has 

elapsed. Trends and changes are already evident, however, for 

some pesticides in selected localities and regions. 

8

This chapter provides selected 

examples of trends and 

illustrates the range of different 

factors that govern trends for 

various pesticides in fish tissue, 

bed sediment, stream water, 

and ground water. 

Long-Term Trends

Photograph used with permission from Larry Mayer
1970s were lower than in 1969, and then continued to decline slowly through the early 1990s. 

Variability in trends in organochlorine pesticides 

during the 1990s, which is evident in figure 8–1, 

probably represents differences among groups of 

NAWQA sites rather than actual trends. 

The observed trends reflect the regulatory history of these three insecticides in the 

United States. Agricultural uses of all three were 

cancelled during the early 1970s, whereas use 

of aldrin, dieldrin, and chlordane was permitted 

for termite control through the late 1980s. The 

declines shown in figure 8–1 are consistent with 

an exponential rate of decline in which concentrations decrease by half within a constant interval of time (half-life), following the elimination 

of use. Nationally, the half-lives in whole fish, 

as estimated either from the NCBP data alone 

or from the combined NCBP and NAWQA data, 

are about 7 years for total DDT, 11–13 years for 

total chlordane, and about 30 years for dieldrin. 

The declines in concentrations of total DDT, 

total chlordane, and dieldrin in whole-fish tissue 

over the past three decades reflect past regulatory 

actions to discontinue their use, yet also illustrate 

that changes can take a long time to occur for 

pesticides with long half-lives.

Data Used to Evaluate Trends in Organochlorine 

Concentrations in Fish Tissue

Although few sites were sampled by both the NCBP and the NAWQA Programs, the sites used in this comparison had similar land uses, and the fishsampling and compositing strategies of the two programs were comparable. 

Of the 117 NCBP sites, most were sampled every 1–3 years during 1969–1986 

(Schmitt and Bunck, 1995). Each of the 228 NAWQA sites was sampled only 

once during 1992–2001, and sites generally were sampled in three groups 

corresponding to the rotational investigations of NAWQA Study Units (see 

Chapter 3). NAWQA sites plotted as 1992 in figure 8–1 actually were sampled 

during the period 1992–1994 (with most sampled during the first year), sites 

plotted as 1995 were sampled during 1995–1997, and sites plotted as 1998 

were sampled during 1998–2001. Because NAWQA sampled three groups of 

sites in three different time periods, the variability in NAWQA results includes 

differences among sites as well as differences over time.

There were also some differences in analytical methods between NAWQA 

and NCBP. The NCBP measured only concentrations of p,p’ isomers of DDT, 

DDD, and DDE in whole fish. For consistency, therefore, NAWQA data for the 

o,p’ isomers of DDT, DDD, and DDE were not included when computing total 

DDT for evaluating trends in whole fish (fig. 8–1). On average, the p,p’ isomers 

of DDT, DDD, and DDE made up 99 percent of total DDT (the sum of o,p’ and 

p,p’ isomers DDT, DDD, and DDE) in whole fish. 

Organochlorine Pesticide Compounds 

in Fish

One of the most striking trends evident 

from historical data and more recent NAWQA 

findings is a national reduction in concentrations 

of organochlorine insecticides in fish tissue, as 

illustrated by concentrations of total DDT, total 

chlordane, and dieldrin measured by NAWQA 

and the U.S. Fish and Wildlife Service. NAWQA 

data from 1992 to 2001 for whole fish in streams 

draining watersheds with mixed land use were 

combined with 1969–1986 data from the U.S. 

Fish and Wildlife Service’s National Contaminant Biomonitoring Program (NCBP), which 

sampled mainly large streams in watersheds with 

mixed land use.

The median and 90th-percentile concentrations of total DDT in whole fish declined markedly from 1969 to about the mid-1970s, with less 

dramatic declines through the 1990s (fig. 8–1). 

Concentrations of total chlordane in fish, for 

which consistent data were not available until 

1978, declined similarly during the 1980s and 

appeared to level off during the 1990s. For dieldrin, the median and particularly the 90th percentile concentrations varied substantially during the 

early 1970s, but concentrations during the late 

128 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 8

Figure 8–1. Concentrations of total DDT, total chlordane, and 

dieldrin in whole fish collected from streams draining watersheds 

with mixed land use throughout the United States have declined 

over the last 20 to 30 years. The declines followed discontinuation 

of their uses during the 1970s (agricultural uses of all three) and 

1980s (use of aldrin, dieldrin, and chlordane for termite control). 

Despite the national decline in concentrations, these persistent 

compounds still are frequently detected in fish. (Data from 1969 

to 1986 are from the U.S. Fish and Wildlife Service, Schmitt and 

Bunck, 1995; and data shown from 1992 to 1998 are from NAWQA. 

All concentrations are for wet weight of fish tissue.)

500

0

1,000

1,500

2,000

2,500

3,000

3,500

4,000

1969 1974 1979 1984 1989 1994 1999

1969 1974 1979 1984 1989

1969 1974 1979 1984 1989 1994 1999

1994 1999

Concentration,

in micrograms per kilogram

90th percentile

Median

100

0

200

300

400

500

600

Concentration,

in micrograms per kilogram

Concentration,

in micrograms per kilogram

0

50

100

150

200

250

300

350

Total DDT

Total Chlordane

Dieldrin

Trends in organochlorine pesticide compounds 

in whole-fish tissue

Long-Term Trends 129
Figure 8–2. Concentrations of total DDT 

declined in most sediment cores collected 

from 41 lakes in 16 States, most of which were 

located in urban areas. The sediment cores 

were analyzed to track historical changes 

from about 1970 through 2000. The downward 

trends are consistent with historical changes in 

DDT use. Upward and downward trends in the 

concentrations of chlordane, however, were more 

evenly distributed, reflecting its continued use to 

control termites until at least 1990.

In addition to the studies of organochlorine pesticides in fish and 

bed sediment from streams, NAWQA also assessed long-term trends 

in the concentrations of these compounds through the analysis of 

sediment cores from lakes and reservoirs, which “record” a history 

of contaminant concentrations. As soils in a watershed erode, they 

are deposited as sediment in layers on the bottom of downstream 

lakes and reservoirs, along with organic particles from aquatic plants 

and animals. Age-dated sediment cores that penetrate these layered 

deposits can be used to track trends in total DDT and total chlordane, 

as well as other sediment-associated contaminants that are relatively 

stable over time. 

Sediment cores from 41 lakes and reservoirs in 16 States—

collectively referred to as lakes for the purposes of this report—were 

analyzed by Van Metre and Mahler (2005). The study included 31 lakes 

in urban settings, 7 lakes in undeveloped settings, and 3 additional 

lakes in watersheds dominated by agriculture. Urban lakes were 

selected to represent watersheds with primarily residential and commercial land uses; only a few of the sites are known to be influenced 

by significant point-source discharges. The statistical significance of 

trends (assessed at the 90-percent confidence level) was determined 

using the Kendall’s tau test for trends in concentrations with depth in 

the core (which is directly related to sediment age). Trends were also 

evaluated by comparing mean decadal concentrations for 1965–1975 

with those for the period from 1990 to approximately 2000 (the top of 

the core).

Concentrations of total DDT (defined by Van Metre and Mahler 

(2005) as the sum of the concentrations of p,p’-DDT, p,p’-DDE, and 

p,p’-DDD) declined significantly since about 1970 in all three of the 

agricultural lakes, 58 percent of the urban lakes, and 43 percent of the 

lakes in undeveloped watersheds (fig. 8–2). No lake had a significant 

upward trend in total DDT. Decadal mean concentrations declined in 

most lakes from 1965–1975 to 1990–2000, including the lakes without 

statistically significant trends within the core samples. Overall, the 

mean total DDT concentration during 1990–2000 was lower than the 

130 Pesticides in the Nation’s Streams and Ground Water, 1992–2001

Trends in Total DDT and Chlordane in Lake Sediments 

Peter C. Van Metre, U.S. Geological Survey

Analyses of sediment cores from lakes 

were used to reconstruct historical 

trends in DDT, chlordane, and other 

contaminants.

Total DDT

Chlordane

Increasing trend

No trend

Decreasing trend

Insufficient detections

Urban (only trend result

symbol is shown)

Undeveloped

Agricultural

Trend result Watershed land use

Trends in sediment cores at lake sites

130 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Figure 8–3. Decreases in the concentrations 

of DDT typically followed an exponential 

curve after uses began to decline during the 

1960s and were cancelled in the early 1970s, 

as shown in sediment cores collected from 

White Rock Lake, Texas, and Lake Blackshear, 

Georgia. The rate of change indicates 

that an additional 50-percent reduction in 

concentrations of total DDT is likely to occur by 

approximately 2015.

mean 1965–1975 concentration in 90 percent of the lakes (37 out of 

41). The median change in total DDT concentration was a decrease of 

68 percent.

These results are consistent with the historical use and regulatory 

history of DDT, as well as with trends in total DDT concentrations in 

whole fish (fig. 8–1). As was observed in fish tissue, the decreases 

in DDT and other persistent hydrophobic contaminants typically followed an exponential curve after use was discontinued (Van Metre 

and others, 1998), with a steep initial drop followed by a gradual 

slowing of the change (fig. 8–3). The resulting half-lives for total DDT 

concentrations in lake sediment range between about 10 and 15 years 

(Van Metre and others, 1998; Van Metre and Mahler, 2005). Applied in 

this way, the half-life does not represent a single specific process (for 

example, chemical degradation), but is a simple measure of the rate 

of change in lake sediment concentrations over time as a result of a 

combination of reduced input, chemical transformations, and dilution. The rate of change observed for total DDT provides an indication 

of what might be expected in the future—an additional 50-percent 

reduction from present concentrations of these compounds during 

the next 10 to 15 years.

Trends in total (technical) chlordane (estimated by Van Metre and 

Mahler [2005] from the concentrations of cis-chlordane, trans-chlordane, and trans-nonachlor) in the sediment cores were more variable 

than those for total DDT, with upward and downward trends split 

evenly, and most showing either no significant trend or insufficient 

data (fig. 8–2). Sixteen percent of urban lakes showed significant 

downward trends, 19 percent showed upward trends, 42 percent 

showed no trend, and 22 percent could not be tested for trends 

because of insufficient detections. Lakes in undeveloped watersheds 

either showed no trend (29 percent), or could not be tested for trends 

because of insufficient detections (71 percent). Only one of the three 

agricultural lakes could be tested for a trend and it was significantly 

downward.

As with total DDT, these results for total chlordane are generally 

consistent with its historical use and regulation. Chlordane use in 

agriculture, which was primarily for corn, was discontinued in 1978; 

however, chlordane use for termite control exceeded its use in agriculture (Andrilenas, 1974; Esworthy, 1987) and continued until 1988 

or later (USEPA, 2004g). In addition, use of existing chlordane stocks 

by homeowners was permitted after 1988 and was common in a 1990 

survey (Whitmore and others, 1992). Therefore, it is not surprising that 

most urban lakes did not show significant downward trends from 1970 

to 2000. This result for urban lakes, however, contrasts with the clear 

decline in total chlordane in fish from watersheds with mixed land 

uses (fig. 8–1). The difference may be caused by the contrasts in land 

use—the fish data from large watersheds with mixed land use may be 

influenced more by agricultural lands, where chlordane use stopped 

in 1978. Another possible explanation for the apparent upward trends 

in some of the urban lake cores is that chemical degradation of one 

or more of the chlordane-derived compounds could be occurring in 

some of the deeper core samples, thus making it appear that concentrations have increased over time (Van Metre and Mahler, 2005). 

The cancellation of DDT and chlordane uses has generally 

resulted in decreased contaminant levels in samples of sediment and 

fish tissue from lakes. However, the continuing high levels of chlordane in urban areas, the slow rate of decreasing trends for DDT, and 

the continuing concern for human exposure from consumption of fish 

and shellfish (USEPA, 2004h, 2005g) indicate that these organochlorine pesticides will remain a concern for many years to come. 

1950

1960

1970

1980

1990

2000

0 20 40 60

Lake Blackshear 

(agricultural)

White Rock Lake 

(urban)

Concentration,

in micrograms per kilogram

Year

Total DDT in sediment cores

Trends in DDT in an agricultural

and an urban lake

Chapter 8

Long-Term Trends 131
Figure 8–4. Concentrations of herbicides 

measured in the White River (White River Basin) 

during 1992–2001 show the correlation between 

stream concentrations and the regional trends 

in use intensity in Corn Belt States. The most 

dramatic examples are the increase in acetochlor 

concentrations after its introduction in 1994 and 

the decreases in alachlor and cyanazine that 

followed reductions in their use. (Pesticide use data 

are from the National Agricultural Chemical Use 

Database, accessed January 25, 2006 at http://www.

pestmanagement.info/nass/app_usage.cfm.)

Median concentration 

during the 

growing season

 (April through August), 

in micrograms per liter

10

1992 1994 1996 1998 2000

1992 1994 1996 1998 2000

Atrazine

Alachlor

Acetochlor

Cyanazine

Metolachlor

1

0.1

0.01

0.001

0.0001

0.001

0.01

0.1

1 Herbicide Use (IA, IL, IN, MN, NE, OH)

Trends in herbicide use and stream 

concentrations

White River at Hazleton, Indiana

Use intensity, 

in pounds per acre

of total agricultural land

Nondetections are plotted at 0.0005 μg/L

Herbicides in Agricultural Streams of 

the Corn Belt

Concentrations of modern, relatively 

short-lived pesticides in stream water generally 

respond rapidly to changes in use. Concentrations of the most heavily used herbicides in 

streams in the Corn Belt showed both increases 

and decreases on a regional scale from 1992 

to 2001, correlating with changes in use during the same period (see sidebar on p. 133). For 

example, concentrations of atrazine, alachlor, 

acetochlor, cyanazine, and metolachlor in the 

White River—a large stream in Indiana that 

drains an extensive agricultural area dominated 

by corn and soybeans—followed regional trends 

in use (fig. 8–4). Acetochlor concentrations in 

the White River rapidly increased following its 

introduction in 1994, whereas alachlor concentrations decreased to less than one-tenth of its 1994 

concentrations by 2001, as acetochlor replaced 

part of alachlor use (note the logarithmic scale in 

figs. 8–4 and 8–5). Among these five herbicides, 

the concentrations of atrazine changed the least 

through the decade, consistent with its relatively 

stable use during this time. Cyanazine concentrations declined most dramatically, following the 

reductions in its use, which began in the mid1990s. The consistency of these trends in the 

region is illustrated by cyanazine results from 

1996 to 2001 for streams in five different States 

within the Corn Belt (fig. 8–5).

132 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 8

Figure 8–5. The consistency of declines in 

cyanazine concentrations in streams throughout 

the Corn Belt is illustrated by results for streams 

in five different States during 1996–2001. Similar 

consistency was evident for other major herbicides 

as well.

8_05

Median concentration during

 the growing season

 (April through August),

in micrograms per liter

10

1996 1997 1998 1999 2000 2001

Nondetections are plotted at 0.0005 μg/L

White River at Hazelton, IN

Maumee River at Waterville, OH

Illinois River at Valley City, IL

Maple Creek near Nickerson, NE

Iowa River at Wapello, IA

1

0.1

0.01

0.001

0.0001

Cyanazine in Corn-Belt streams

Trends in Use of Herbicides in the Corn Belt 

From 1992 to 2001, there were major changes in the primary herbicides used for corn and soybean production in the Corn Belt States of Iowa, Illinois, Indiana, Minnesota, Nebraska, and Ohio (see fig. 8–4), even 

though the total treated crop acreage remained fairly constant. Changes in these herbicides, which combine to account for more than 40 percent of all national herbicide use, exemplify the shifts in use patterns 

that are typical of pesticides in response to changes in factors such as regulations, monitoring results, 

effectiveness, and cost.

Throughout the 1990s, atrazine was the herbicide used most widely on corn, and the area treated 

each year varied little. In the early 1990s, atrazine use decreased slightly because of reduced application 

rates resulting from regulatory agreements between USEPA and the atrazine manufacturer. This decline, 

however, was soon offset by increased use of atrazine in tank mixes with other herbicides, and total use 

remained near the 1990 level in major corn-producing States. 

Metolachlor and alachlor were the second and third most heavily used herbicides in 1990, but their use 

declined substantially by 2001 because of the introduction of new herbicides. In 1994, acetochlor was conditionally registered for use on corn, with the goal of reducing the use of alachlor and other corn herbicides 

by one-third. By 1997, acetochlor had virtually replaced alachlor use and was rapidly becoming one of the 

most widely used herbicides (note the logarithmic scale in fig. 8–4). Also in 1997, S-metolachlor was conditionally registered for use. S-metolachlor is the more effective form of two different isomers of metolachlor 

(both the R- and S-metolachlor isomers were present in metolachlor products). The introduction of S-metolachlor, which has a 30-percent lower application rate, contributed to the decrease in total metolachlor 

use during the late 1990s. An additional development that probably contributed to the decline in the use of 

metolachlor and other herbicides was the introduction of bioengineered crops that were genetically modified to be resistant to specific herbicides, such as glyphosate. 

The most dramatic decline in herbicide use during the 1990s was for cyanazine. Because of frequent 

detection of cyanazine in surface and ground water, cyanazine manufacturers began to phase out this 

product beginning in 1994. This phase-out, which was completed in 2000, shifted cyanazine from the fourth 

most heavily used herbicide in 1992, to only minor use by 2001.

Long-Term Trends 133
Until about 2001, diazinon was one of the most widely used insecticides in the United States for residential lawn and garden pest control (accounting for almost 70 percent of the 11 million lb of diazinon 

used for all purposes each year), for residential indoor uses (up to 5 

percent of total use), and for agricultural pest control (almost 30 percent of total use). In December 2000, USEPA and diazinon registrants 

agreed to phase out the sale of diazinon for residential uses (both 

outdoor and indoor), as well as for many agricultural uses. As part of 

the agreement, indoor uses of diazinon were terminated and all outdoor nonagricultural uses (principally on residential lawns and gardens) were phased out during 2002–2004. Manufacturing of diazinon 

for application to gardens, lawns, and other turf stopped in June 2003, 

and sales and distribution to retailers ended in August 2003. Retail 

sales ended on December 31, 2004, after which a buy-back program 

helped to remove from the market the remaining diazinon products. 

USEPA and the registrants also agreed to reduce the uses of diazinon 

on agricultural crops by about one-third. By 2005, these combined 

actions eliminated most of the use of diazinon, compared with use in 

2000. 

Analysis of data from seven NAWQA stream sites in the Northeast—five classified as urban streams and two as mixed land 

use—using Seasonal Kendall tests at the 95-percent confidence 

level (Schertz and others, 1991), indicate predominantly downward 

trends in concentrations of diazinon since the reductions in diazinon 

use began in 2000 (fig. 8–6). Specifically, concentrations of diazinon 

decreased by 20 to 41 percent since 1998 at the five sites with statistically significant changes. Concentrations at the two sites with no statistically significant change showed decreases of 13 and 22 percent 

for the same period. Diazinon concentrations observed in one of the 

five urban streams (Accotink Creek, VA) provide an illustration of how 

concentrations have recently declined in some streams—in this case 

by about 39 percent from 1998 to 2004 (fig. 8–7).

-13

-41

-20

-33

-28

-22 -39

-13 Not statistically significant

-20 Statistically significant

Decreasing trends and percentage change

Trends in diazinon concentrations

at selected stream sites

Figure 8–6. Diazinon concentrations 

decreased significantly from 1998 to 2004 

in 5 of 7 urban and mixed-land-use streams 

in the northeastern United States. Trends 

were evaluated using the Seasonal Kendall 

test at the 95-percent confidence level. 

1997 1998 1999 2000 2001 2002

Open circles ( )

denote nondetections

First year of phasing out of

diazinon use on lawns and gardens

Concentration,

in micrograms per liter

0.001

0.01

0.1

1

2003 2004 2005

Diazinon in Accotink Creek, Virginia

Figure 8–7. From 1997 to 2001, the levels and 

ranges of diazinon concentrations in Accotink 

Creek (Potomac River Basin) were relatively 

similar from year to year, but from 2002 to 2004, 

diazinon concentrations generally decreased. 

These decreases correspond to the national 

reduction in total sales and use of diazinon 

through this period, although no specific use data 

were available for the Accotink Creek watershed.

Recent NAWQA Data Show that Diazinon Concentrations in Some Northeast Streams have 

Declined Following Recent Reductions in Use

Patrick J. Phillips, U.S. Geological Survey

Scott W. Ator, U.S. Geological Survey

Elizabeth A. Nystrom, U.S. Geological Survey

134 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Chapter 8

Ground Water

Pesticide concentrations in ground water, 

compared with streams, respond more slowly 

to changes in pesticide use or land-management 

practices, often lagging by years or decades, 

depending on the nature of the flow system and 

the depth and location of wells sampled. During 

the long periods of time that it takes for water to 

move through most ground-water flow systems, 

the types and amounts of pesticides applied at the 

land surface often change. This makes it difficult 

to link the concentrations of pesticides detected 

in specific wells with the locations where the pesticides were used. Evaluation of trends in ground 

water is also made more difficult by the complex 

flow paths along which ground water moves, and 

the resulting uncertainty about where sampled 

water originally entered the ground-water flow 

system. 

For these reasons—as well as a general 

shortage of suitable data—trends in pesticide 

levels in ground water have not been extensively 

characterized. As noted by Barbash and Resek 

(1996), few previous studies have used consistent sampling and analytical methods over 

long enough periods, or developed a sufficient 

understanding of the flow system and age of 

sampled ground water, to reliably evaluate longterm trends in ground-water quality. Although 

NAWQA ground-water studies use consistent 

sampling and analytical methods, NAWQA monitoring has not yet covered a long enough period 

of time in most locations to assess trends. Despite 

these challenges, examples of ground-water trend 

assessments from USGS studies, conducted in 

cooperation with other agencies in Iowa and Florida, illustrate the types of trends that may occur 

over different time scales and demonstrate some 

of the approaches to trend assessment.

Herbicides in Iowa Ground Water

Results from the Iowa Ground Water 

Monitoring Program, a joint study by the Iowa 

Geological Survey, USEPA, and USGS, show 

that herbicide concentrations have increased in 

Iowa ground water with increasing herbicide use 

since the 1950s (Kolpin and others, 2004). Low 

levels of tritium (less than 0.8 tritium units [TU]) 

were used as an indicator of water recharged 

before 1953, which was prior to the onset of 

substantial herbicide use. All but 1 of 42 samples 

with detectable concentrations of herbicides 

or degradates were samples of water that had 

recharged after 1953 (fig. 8–8), whereas more 

than 80 percent of the samples with undetectable 

herbicides or degradates had recharged prior to 

1953. The detection of herbicides in one sample 

with low tritium probably resulted from the 

mixing of younger and older waters in samples 

collected from a municipal supply well. The 

results from this study demonstrate the value of 

information on ground-water recharge dates and 

residence times for the analysis of data for trend 

assessment. Use of estimated recharge dates 

provided the most reliable means available for 

determining that most samples without detections 

were ground water that had recharged before 

the beginning of major herbicide use. In addition, because the correlations between estimated 

recharge date and the occurrence of pesticide 

Figure 8–8. The Iowa Ground-Water 

Monitoring Program showed that 

herbicides or their degradates were 

detected in 41 of 42 municipal supply 

wells in Iowa that tap ground water 

recharged after 1953 (as indicated by 

tritium concentrations greater than 

0.8 tritium units). Conversely, more 

than 80 percent of the samples in 

which herbicide compounds were 

not detected were samples of ground 

water recharged prior to 1953, before 

significant use of herbicides began. 

(Modified from Kolpin and others, 2004.) 0.01

0.1

1

10

100

Recharged after 1953

Recharged before 1953

Tritium concentration, in tritium units

Herbicides

or degradates

detected

Herbicides

or degradates

not detected

Median

75th percentile

25th percentile

Detections of herbicide compounds in ground 

water recharged before and after 1953

Long-Term Trends 135
compounds in ground water were more evident 

when degradates were considered, the study by 

Kolpin and others (2004) demonstrates the value 

of incorporating data on degradate occurrence for 

detecting trends.

Herbicides in Florida Ground Water

A study by the USGS, Florida Department 

of Agriculture and Consumer Services, and the 

Southwest Florida Water Management District 

was undertaken to monitor and assess the quality of shallow ground water in central Florida 

(Choquette and Sepúlveda, 2000). This region is 

dominated by citrus production and is characterized by well-drained sandy soils that are conducive to relatively rapid movement of water and 

pesticides to and within the ground-water flow 

system (Choquette and others, 2003).

The study found that bromacil, a widely 

used herbicide, declined but continued to be 

detected in 25 percent or more of the sampled 

wells for up to 10 years after its use in the 

region’s citrus orchards was prohibited in 1994 

(fig. 8–9). The decline in bromacil detections 

coincided with an increase in detections of 

norflurazon, which began to replace bromacil in 

1994. The frequency of norflurazon detections, 

however, did not exceed that of bromacil until the 

year 2000, about 6 years after the use of bromacil 

was discontinued (Choquette and others, 2005).

Although figure 8–9 indicates that the 

overall frequency of norflurazon detection in 

ground water increased from 1993 to 2004 within 

the area studied by Choquette and others (2005), 

concentrations of norflurazon showed considerable seasonal and annual variability in the 

individual wells sampled, as well as variability 

among different wells (fig. 8–10). These variations were associated with differences among the 

wells in the age of the ground water, the depth to 

water, the depths of the sampled zone below the 

water table, and the thickness of the aquifer zone 

sampled. The highest and most variable concentrations (wells 1–3) occurred where depths to the 

water table were relatively shallow and in wells 

that sampled water closest to the water table. The 

lower and less variable concentrations occurred 

in the deeper wells (wells 4–8) with long 

screened intervals. These observations are consistent with results from previous studies, indicating 

that the temporal variability in pesticide concentrations generally tends to diminish with increasing well depth (Barbash and Resek, 1996).

Figure 8–9. In central Florida, applications of bromacil 

in citrus orchards were discontinued in 1994, yet it 

was still detected at or above 2 μg/L in 25 percent of 

the sampled wells 10 years later. The frequency of 

detecting norflurazon, which began to replace bromacil 

on citrus in 1994, did not exceed that of bromacil for 

about 6 years. (Modified from Choquette and others, 

2005.)

Figure 8–10. Although concentrations of norflurazon 

increased from 2000 to 2004, there was substantial 

seasonal and annual variability. Variability among wells 

was associated with depth to water, depth of the well 

screen below the water table, the length of the well 

screen, and the thickness of the aquifer zone sampled. 

(Modified from Choquette and others, 2005.)

1992 1994 1996 1998

Year

2000 2002 2004

25

0

50

75

100

Bromacil

Norflurazon

Trends in herbicide concentrations in ground

water beneath citrus orchards in Florida

Frequency of detection, in percent

8_09

10

0

20

30

40

1

2

3

4

5

6

7

8

Well number

Trends in norflurazon concentrations in individual

wells in citrus orchards in Florida

Concentration, in micrograms per liter

April

July

October

January

April

July

October

January

April

July

October

January

April

July

October

January

April

2000 2001 2002 2003 2004

136 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
References Cited 137

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USEPA (U.S. Environmental Protection Agency), 2000b, 

Environmental risk assessment for diazinon, revised October 2000: U.S. Environmental Protection Agency, accessed 

September 28, 2005, at http://www.epa.gov/pesticides/op/

diazinon/risk_oct2000.pdf.

USEPA (U.S. Environmental Protection Agency), 2000c, 

Supplementary guidance for conducting health risk assessment of chemical mixtures: U.S. Environmental Protection Agency Risk Assessment Forum EPA/630/R-00/002, 

accessed December 1, 2005, at http://www.epa.gov/NCEA/

raf/chem_mix.htm.

USEPA (U.S. Environmental Protection Agency), 2003a, 

Interim reregistration elegibility decision for atrazine—Case 

No. 0062: U.S. Environmental Protection Agency, accessed 

January 9, 2006, at http://www.epa.gov/oppsrrd1/REDs/

atrazine_ired.pdf.

USEPA (U.S. Environmental Protection Agency), 2003b, 

Atrazine MOA Ecological Subgroup—recommendations 

for aquatic community level of concern (LOC) and method 

to apply LOC(s) to monitoring data, final report, October 

22, 2003, U.S. Environmental Protection Agency Electronic 

docket, Public Docket Number OPP-2003-0367-0007, 

accessed January 24, 2006, at http://www.regulations.gov/

fdmspublic-rel11/component/main. 

USEPA (U.S. Environmental Protection Agency), 2003c, 

Procedures for the derivation of equilibrium partitioning 

sediment benchmarks (ESBs) for the protection of benthic organisms— dieldrin: U.S. Environmental Protection 

Agency, Office of Research and Development EPA-600-

R-02-010, accessed December 1, 2005, at http://www.epa.

gov/nheerl/publications/files/dieldrin.pdf. 

USEPA (U.S. Environmental Protection Agency), 2003d, 

Procedures for the derivation of equilibrium partitioning 

sediment benchmarks (ESBs) for the protection of benthic 

organisms—endrin: U.S. Environmental Protection Agency, 

Office of Research and Development EPA-600-R-02-009, 

accessed December 1, 2005, at http://www.epa.gov/nheerl/

publications/files/endrin.pdf.

USEPA (U.S. Environmental Protection Agency), 2003e, 

Procedures for the derivation of equilibrium partitioning 

sediment benchmarks (ESBs) for the protection of benthic 

organisms—nonionics compendium, Draft: U.S. Environmental Protection Agency, Office of Research and Development EPA-822-R-02-016.

USEPA (U.S. Environmental Protection Agency), 2003f, 

Framework for cumulative risk assessment: U.S. Environmental Protection Agency Risk Assessment Forum 

EPA/630/P-02/001F, 129 p.

USEPA (U.S. Environmental Protection Agency), 2004a, 40 

CFR Part 158 Data Requirement Tables: U.S. Environmental Protection Agency.

142    Pesticides in the Nation’s Streams and Ground Water, 1992–2001
USEPA (U.S. Environmental Protection Agency), 2004b, An 

examination of EPA risk assessment principles and practices—staff paper prepared for the U.S. Environmental Protection Agency by members of the Risk Assessment Task 

Force: U.S. Environmental Protection Agency, Office of the 

Science Advisor EPA/100/B-04/001, 193 p. 

USEPA (U.S. Environmental Protection Agency), 2004c, 

2004 Edition of the drinking water standards and health 

advisories: U.S. Environmental Protection Agency, Office 

of Water EPA-822-R-04-005, accessed December 1, 2005, 

at http://www.epa.gov/waterscience/drinking/standards/

dwstandards.pdf.

USEPA (U.S. Environmental Protection Agency), 2004d, 

National recommended water quality criteria: U.S. Environmental Protection Agency, Office of Water and Office of 

Science and Technology, 23 p., accessed December 1, 2005, 

at http://www.epa.gov/waterscience/criteria/nrwqc-2004.

pdf.

USEPA (U.S. Environmental Protection Agency), 2004e, 

Interim reregistration eligibility decision diazinon: U.S. 

Environmental Protection Agency, Prevention, Pesticides 

and Toxic Substances, EPA 738-R-04-006 (May 2004), 

accessed September 21, 2005, at http://cfpub.epa.gov/

oppref/rereg/status.cfm?show=rereg.

USEPA (U.S. Environmental Protection Agency), 2004f, 

National whole effluent toxicity (WET) implementation 

guidance under the NPDES program (draft): U.S. Environmental Protection Agency, Office of Wastewater Management EPA 832-B-04-003, accessed December 1, 2005, at 

http://www.epa.gov/npdes/pubs/wet_draft_guidance.pdf.

USEPA (U.S. Environmental Protection Agency), 2004g, Persistent bioaccumulative and toxic chemical program—priority PBT profiles: U.S. Environmental Protection Agency, 

accessed January 10, 2006, at http://www.epa.gov/pbt/pubs/

cheminfo.htm.

USEPA (U.S. Environmental Protection Agency), 2004h, The 

incidence and severity of sediment contamination in surface 

waters of the United States, national sediment quality 

survey (2nd ed.): U.S. Environmental Protection Agency, 

Office of Science and Technology EPA 823-R-04-007, 

280 p.

USEPA (U.S. Environmental Protection Agency), 2005a, 

Technical overview of ecological risk assessment: U.S. 

Environmental Protection Agency, accessed January 27, 

2006 at http://www.epa.gov/oppefed1/ecorisk_ders/. 

USEPA (U.S. Environmental Protection Agency), 2005b, 

Drinking water contaminant candidate list 2—final notice: 

Federal Register, v. 70, no. 36, p. 9071-9077, February 24, 

2005; accessed December 1, 2005, at http://www.epa.gov/

fedrgstr/EPA-WATER/2005/February/Day-24/w3527.htm.

USEPA (U.S. Environmental Protection Agency), 2005c, 

Unregulated contaminant monitoring regulation (UCMR) 

for Public Water Systems Revisions—proposed rule: 

Federal Register, v. 70, no. 161, p. 49093–49138, accessed 

December 1, 2005, at http://www.epa.gov/fedrgstr/EPAWATER/2005/August/Day-22/w16385.htm.

USEPA (U.S. Environmental Protection Agency), 2005d, 

Drinking water health advisories: U.S. Environmental 

Protection Agency, accessed November 14, 2005, at http://

www.epa.gov/waterscience/criteria/drinking/.

USEPA (U.S. Environmental Protection Agency), 2005e, 

Integrated Risk Information System (IRIS) database for 

risk management: U.S. Environmental Protection Agency, 

Office of Research and Development, National Center for 

Environmental Assessment, accessed December 1, 2005, at 

http://www.epa.gov/iris.

USEPA (U.S. Environmental Protection Agency), 2005f, 2004 

National listing of fish advisories: U.S. Environmental 

Protection Agency, Office of Water, Fact Sheet EPA-823-

F-05-004, 6 p., accessed December 2, 2005, at http://epa.

gov/waterscience/fish/advisories/fs2004.pdf. 

USEPA (U.S. Environmental Protection Agency), 2005g, 

The national listing of fish advisories database, release of 

2004 data: U.S. Environmental Protection Agency, Office 

of Water, accessed November 9, 2005, at http://epa.gov/

waterscience/fish/advisories/.

USEPA (U.S. Environmental Protection Agency), 2005h, 

Technical overview of ecological risk assessment, risk 

characterization: U.S. Environmental Protection Agency, 

accessed November 16, 2005, at http://www.epa.gov/

oppefed1/ecorisk_ders/toera_risk.htm. 

USEPA (U.S. Environmental Protection Agency), 2005i, 

Technical overview of ecological risk assessment, analysis 

phase— ecological effects characterization: U.S. Environmental Protection Agency, accessed January 9, 2006, at 

http://www.epa.gov/oppefed1/ecorisk_ders/toera_analysis_

eco.htm.

USEPA (U.S. Environmental Protection Agency), 2005j, 

Proposed revisions to the aquatic life criteria guidelines—

charge to the panel: U.S. Environmental Protection Agency, 

accessed November 17, 2005, at http://www.epa.gov/

science1/pdf/aquatic_life_criteria_guidelines_consultative_

panel_charge.pdf.

U.S. Geological Survey, 1999, The quality of our Nation’s 

waters—nutrients and pesticides: U.S. Geological Survey 

Circular 1225, 82 p.

References Cited 143
Van Metre, P.C., Wilson, J.T., Callender, Edward, and Fuller, 

C.C., 1998, Similar rates of decrease of persistent, hydrophobic and particle-reactive contaminants in riverine systems: Environmental Science and Technology, v. 32, no. 21, 

p. 3312–3317.

Van Metre, P.C., and Mahler, B.J., 2005, Trends in hydrophobic organic contaminants in urban and reference lake sediments across the United States, 1970–2001: Environmental 

Science and Technology, v. 39, no. 15, p. 5567–5574.

Vogelmann, J.E., Howard, S.M., Yang, L., Larson, C.R., 

Wylie, B.K., and VanDriel, Nick, 2001, Completion of the 

1990’s national land cover dataset for the conterminous 

United States from Landsat Thematic Mapper data and 

ancillary data sources: Photogrammetric Engineering and 

Remote Sensing, v. 67, p. 650–662.

Whitmore, R.W., Kelly, J.E., and Reading, P.L., 1992, Executive summary, results, and recommendations, Volume 1 of 

National home and garden pesticide use survey, final report: 

U.S. Environmental Protection Agency, Research Triangle 

Institute RTI/5100/17-01F, 140 p.

 

144 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Glossary 145

Glossary

(Terms in definitions that are defined elsewhere in the Glossary are shown in boldface within the definition. Most definitions are described as they apply to pesticides.)

A

10–6 cancer risk concentration (CRC) The concentration of 

a chemical in drinking water that corresponds to an excess 

estimated lifetime cancer risk (in addition to cancer risk from 

other causes) of 1 in 1,000,000.

Acetanilide herbicides A class of pesticides derived from 

N-acetylaniline and used primarily for weed control in corn, 

soybeans, and sorghum.

Active ingredient The chemical component of a pesticide 

product that kills or otherwise controls the target organism(s). 

Acute effects Rapid physiological response of an organism (such as death or immobility) resulting from relatively 

short-term exposure to elevated concentrations of one or more 

chemicals or other changes in biological, chemical, or physical 

conditions in the environment.

Adjuvants Chemicals included in a pesticide product to 

facilitate the application of the product or to enhance the effectiveness of the active ingredient. Adjuvants are often referred 

to as “inert ingredients.”

Agricultural management practices Methods used as part 

of crop cultivation and livestock husbandry (such as irrigation, fertilization, or integrated pest management) to maximize 

product yields, control soil erosion, maintain soil quality, and 

(or) minimize any adverse effects on water quality or ecosystem health.

Agricultural stream A stream draining a watershed with 

more than 50 percent agricultural land (cropland or pasture) 

and 5 percent or less of urban land.

Ambient water-quality criteria (AWQC) Guidelines issued 

by USEPA for pollutants designated as toxic under the Clean 

Water Act and that may provide the basis for state standards. 

There are two types of these guidelines—those for the protection of human health and those for the protection of aquatic 

organisms. Aquatic-life criteria may be acute (established for 

short-term exposure) or chronic (for long-term exposure). 

Aquatic-life benchmark A threshold value above which the 

concentration of a chemical in water or bed sediment may 

have adverse effects on aquatic organisms. Benchmarks for 

water are established to address either acute (short-term) or 

chronic (long-term) exposures.

Aquifer A geologic formation, group of formations, or part 

of a formation that contains a sufficient amount of saturated 

permeable material (for example, soil, sand, gravel and (or) 

rock) to yield significant quantities of water to wells and 

springs.

B

Baseflow Hydrologic regime in streams, following extended 

periods of minimal precipitation, during which streamflow is 

derived primarily from ground-water discharge.

Bed sediment Sediment particles, including eroded soil and 

organic matter, deposited at the bottom of a stream or other 

surface-water body. 

Benchmark See water-quality benchmark.

Benthic Living on or close to the bottom of a stream, lake, 

or sea.

Bioaccumulative The tendency of a chemical compound to 

be taken up and retained by organisms from all sources in their 

environment, such as diet, sediment, soil, or water.

Bioassay The quantitative measurement, under standardized conditions, of the biological effects of a substance on an 

organism or part of an organism.

C

Carbamate insecticides A class of pesticides consisting of 

various esters of carbamic acid. Like the organophosphate 

insecticides, they are inhibitors of cholinesterase—the enzyme 

required for nerve function in the animal body—and are used 

to kill or control insects in a variety of agricultural and nonagricultural settings.

Chlordane group The set of five compounds whose concentrations are summed to compute the concentration of total 

chlordane, including the cis and trans isomers of chlordane 

and nonachlor, and the chlordane degradate, oxychlordane.

Chronic effects Physiological responses of an organism 

(such as death, impaired reproduction, or changes in organ 

function) resulting from long-term exposure to one or more 

chemicals or other changes in biological, chemical, or physical 

conditions in the environment.

Common detection level A single concentration threshold, 

used for assessing the presence or absence of each one of a 

group of compounds within a sample or set of samples on an 

equal basis. Use of this threshold avoids biases in detection 

frequencies caused by varying analytical sensitivities to different compounds—it is also sometimes referred to as a “common assessment level.”

Confidence level The probability threshold used to decide 

whether a particular observation or result of a statistical test 

was likely to have arisen solely by chance.

Conventional pesticides Compounds that are commonly 

used to kill or control unwanted organisms in either agricultural or nonagricultural settings. Such chemicals include 

herbicides, insecticides, fumigants, fungicides, and many 

other types of biocidal compounds, but exclude several other 

types, such as antifouling agents, disinfectants, and wood 

preservatives.

Corn Belt The area of the Great Plains and the Midwest 

where corn and soybeans are the principal crops. It generally 
includes Illinois, Indiana, Iowa, Nebraska, and Ohio, as well 

as parts of adjoining states.

Crop-group setting A classification of agricultural land that 

is based on the dominant presence of one, two, or three specific crops (such as “rice,” or “corn and soybeans”), as derived 

from the classification system described by Gilliom and Thelin 

(1997).

D

DDT group Six compounds derived from the parent pesticides DDT and DDD whose concentrations are summed to 

compute the concentration of total DDT, specifically the p,p'

and o,p' isomers of DDT, DDD, and DDE.

Deethylatrazine-to-atrazine ratio The ratio of concentrations 

of deethylatrazine to atrazine in a particular environmental 

medium (usually water), used to track the transformation 

of atrazine to one of its principal degradates over time or 

distance.

Degradate A compound produced from the transformation 

of a parent pesticide or another degradate through either 

abiotic or biotic processes.

Diffusion The movement of chemicals (in either the gas, 

liquid, or solid phase) from regions of higher concentration to 

those of lower concentration.

Domestic well A privately owned well that usually serves 

one home and supplies water for human consumption and 

other homeowner uses. 

E

EC50 In a toxicity test, the “50 percent effect concentration”—that is, the concentration of a chemical at which a 

specified effect is observed in half of the test organisms within 

a specified period of time (typically 48 hours). 

Endocrine disruptor A chemical that interferes with the 

endocrine system in an organism by mimicking a natural 

hormone, blocking the effects of a hormone on certain receptors, or causing the overproduction or underproduction of 

hormones.

Endocrine system A biochemical regulatory system in 

mammals, birds, fish, and other organisms that is comprised 

of hormones (which act as chemical messengers), glands that 

produce hormones, and receptors in various organs and tissues 

that recognize and respond to the hormones. The system regulates a wide variety of physiological processes in the body, 

including the development of the brain and nervous system, 

the growth and function of the reproductive system, metabolic 

activity, and blood sugar levels. 

Environmental medium Any natural solid, liquid, or gas in 

the environment—such as ground water, stream water, bed 

sediment, or biological tissues.

Explanatory variable A parameter (for example, chemical 

use, population, or soil permeability) whose value is used in 

regression and other statistical models to evaluate and estimate 

the magnitude of another parameter (the response variable).

F

Fish-consumption advisory A recommendation issued by a 

public agency that people limit or avoid consumption of certain fish species caught from particular water bodies because 

of contamination of fish with bacterial or bioaccumulative

pollutants.

Flow path The route or pathway followed by water flowing 

through the hydrologic system. Usually refers to subsurface

flow.

Fumigant A compound or mixture of compounds that produces a gas, vapor, fumes, or smoke intended to destroy, repel, 

or control organisms such as insects, bacteria, or rodents.

Fungicides Pesticides that are used to kill unwanted fungi.

G

Glacial till Poorly sorted unconsolidated geologic material 

deposited by glaciers and generally having low permeability, 

unless fractures or other interconnected openings for flow are 

present.

Ground-water recharge Water that reaches ground water 

by infiltration of precipitation or irrigation water through the 

unsaturated zone or by seepage of water from surface-water 

bodies, such as streams and lakes.

Guideline A threshold value for the maximum acceptable concentration of a pesticide or other contaminant in a 

given environmental medium, specified for the protection 

of humans, aquatic life, or wildlife. Guidelines are issued for 

advisory purposes and are not legally enforceable. 

H

Half-life The time required for the concentration of a compound in a given environmental medium to be reduced to 

half of its original value by one or more processes, such as 

degradation or transport into another environmental medium.

Health advisory An estimate of acceptable drinking-water 

concentrations for a chemical substance, established by 

USEPA on the basis of health effects information. Although it 

is not a legally enforceable federal standard, it provides technical guidance to assist Federal, State, and local officials.

Henry’s law constant (KH) A measure of the partitioning of a 

compound between an aqueous solution and a gas with which 

it is in contact, quantified as the ratio between the concentrations of the compound in the gas phase and in the aqueous 

solution at equilibrium.

Herbicides Pesticides that are used to kill unwanted plants.

Human-health benchmark A threshold value above which 

the concentration of a chemical in water may have adverse 

effects on humans if the water is used as drinking water without treatment or other measures to lower the concentration.

146 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Hydraulic conductivity The rate at which a porous medium 

transmits water.

Hydrogeology The geologic and hydrologic features that 

control the movement of water, solutes, and small particles 

through the subsurface.

Hydrologic system The assemblage of pathways by which 

water travels as it circulates beneath, at, and above the Earth’s 

surface through various processes such as precipitation, runoff, 

evaporation, infiltration, transpiration, and ground-water 

discharge.

Hydrophilic The tendency of a compound to favor dissolution in, or association with water, rather than organic matter. 

Often used to refer to compounds with comparatively low K

oc

values.

Hydrophobic The tendency of a compound to favor sorption

to, or association with organic matter, rather than dissolution 

in water. Often used to refer to compounds with comparatively 

high K

oc

 values.

I

Immobile zones Regions within the subsurface through 

which water and solutes move relatively slowly, if at all. (Contrast with mobile zones.)

Inert ingredients See adjuvants.

Insecticides Pesticides that are used to kill unwanted 

insects.

Isomers Compounds with identical chemical composition 

but with slightly different structures (arrangement of atoms). 

Examples include o,p'- and p,p'-DDT; and cis- and transchlordane.

K

K

oc

See soil organic carbon-water partition coefficient.

KH See Henry’s law constant.

Kendall’s tau test A nonparametric statistical test used to 

determine whether a particular trend in magnitude is significant at a specified confidence level.

L

LC50 In a toxicity test, the “50 percent lethal concentration”—

that is, the concentration of a chemical at which 50 percent of 

test organisms die within a specified period of time (typically 

48 or 96 hours).

Lifetime Health Advisory (HA-L) The concentration of a 

chemical in drinking water that is not expected to cause any 

adverse noncarcinogenic effects in humans over a lifetime of 

exposure (70 years). This parameter is not a legally enforceable federal standard, but provides technical guidance to assist 

Federal, State, Tribal, and local officials.

Linear regression A statistical method for analyzing and 

estimating the magnitude of a response variable as a function 

of one or more explanatory variables.

Lipid Any one of a diverse group of hydrophobic organic 

compounds produced and stored by living organisms and that 

contain long hydrocarbon chains or rings. Examples include 

fats, oils, waxes, steroids, and carotenoids.

M

Major aquifer A regionally extensive subsurface geologic 

formation or group of formations that is used, or has the potential to be used, as a significant ground-water resource. 

Major aquifer studies NAWQA investigations involving the 

sampling of 20 to 30 domestic and (or) public-supply wells

that withdraw water from major aquifers. 

Manufacturing by-products Compounds used for, or generated during, the production of a particular chemical (such as a 

pesticide active ingredient) that may be present in the commercial formulation itself, especially those such as technical 

mixtures that are less highly purified.

Maximum Contaminant Level (MCL) A drinking-water 

standard that is legally enforceable and that sets the highest 

permissible concentration of a specific compound in water that 

is delivered to any user of a public water system. In this report, 

only values established by the USEPA are used.

Mixed-land-use streams Streams draining watersheds in 

which no single type of land use (agricultural, urban, or undeveloped) predominates. These include all streams not meeting the specific land-use criteria for agricultural, urban, or

undeveloped streams.

Mixture A combination of two or more compounds detected 

in the same environmental sample.

Mobile zones Regions within the subsurface through which 

water flows more rapidly than in other locations, often consisting of worm holes, cracks, fractures, and other highly conductive channels. Also referred to as zones of “preferential flow.” 

(Contrast with immobile zones.)

Mobility The speed or ease with which a compound moves 

through the hydrologic system relative to the rate of water 

flow— mobility generally increases with decreasing K

oc

values.

N

No-effect level In a toxicity study, the highest concentration 

or dose that was observed to have a negligible impact on the 

health of the test organisms.

Nonpoint sources Contaminant releases that are diffuse and 

widely dispersed, such as agricultural runoff or atmospheric 

deposition.

Nontarget organisms Organisms other than those that a pesticide active ingredient is applied to control or kill.

Glossary 147
O

Observation well A well designed for measuring water 

levels and testing ground-water quality, and generally not used 

as a source for drinking water. Also referred to as a “monitoring well.”

Organochlorine pesticide compounds A class of synthetic 

organic chemicals (mostly insecticides) with hydrocarbon 

structures containing one or more chlorine substituents, and 

that includes manufacturing by-products and degradates, in 

addition to active ingredients.

Organophosphate insecticides A group of pesticides, consisting of various derivatives of phosphoric, phosphorothioic, 

or phosphorodithioic acids, in some cases with a nitrogen, 

fluorine, methyl, or cyano group substituting for one or more 

of the phosphate oxygens. Like the carbamate insecticides, 

they are inhibitors of cholinesterase—the enzyme required 

for nerve function in the animal body—and are used to kill or 

control insects in a variety of agricultural and nonagricultural 

settings.

P

Parent pesticide The form of an active ingredient as it is 

released into the environment. 

Partitioning The processes by which a compound becomes 

distributed among different environmental media. Such 

processes include sorption, volatilization, dissolution, and 

biological uptake.

Permeability A measure of the relative ease with which a 

porous medium can transmit a fluid. 

Persistence The tendency of a compound to remain in its 

original form, rather than undergo transformation, in the 

environment.

Pesticide A chemical applied to crops, rights-of-way, lawns, 

residences, golf courses, or other settings to kill or control 

weeds, insects, fungi, nematodes, rodents, or other unwanted 

organisms. 

Pesticide compounds A term used to refer collectively to 

parent pesticides, their degradates and, where applicable, 

their manufacturing by-products.

Point source A specific location at which one or more 

contaminants are known to be released into the hydrologic 

system.

Public-supply well A privately or publicly owned well 

that provides water for public use to: (1) community water 

systems, (2) transient non-community water systems, such as 

campgrounds, or (3) non-transient, non-community systems, 

such as schools.

R 

Rainfall erosivity A parameter in the Universal Soil Loss 

Equation that quantifies the effects of rainfall on soil loss 

within a particular area, and accounts for both the energy and 

intensity of rainstorms, averaged over a specified number of 

years. Also referred to as the “R factor.”

Residence time The amount of time that a solute, particle, 

organism, or other entity spends within a given environmental 

medium.

Response variable The dependent parameter (for example, 

chemical concentration) whose magnitude is estimated from 

quantitative relations with other, independent parameters 

(explanatory variables) using statistical relations such as 

regression models.

Rill irrigation Water management method that employs a 

series of parallel surface ditches to distribute water to crops.

S

Saturated zone The region in the subsurface in which all 

the interstices or voids are filled with water under a pressure 

exceeding that of the atmosphere.

Seasonal Kendall test A statistical method that corrects for 

possible seasonal patterns in a given variable to detect temporal trends in the parameter (or lack thereof) over a period of 

years.

Seasonal pulse Temporary increase in the concentration 

of one or more compounds in surface water or ground water 

that commonly occurs at a particular time of the year—for 

example, the substantial increases in the concentrations of 

corn herbicides typically observed in streams of the Corn 

Belt in the spring.

Simulation model A mathematical model used to predict 

the combined effects and (or) consequences of one or more 

processes of interest by reproducing these effects using mathematical relations and (or) numerical techniques, typically 

through the use of computer programs.

Soil erodibility A parameter in the Universal Soil Loss 

Equation that quantifies the ease with which a given soil may 

be carried away by water and which is based on a number of 

soil characteristics such as soil texture (that is, the percentages 

of different size fractions such as sand, silt, and clay), organicmatter content, permeability, and structure. Also referred to as 

the “K factor.”

Soil organic carbon-water partition coefficient (Koc) A 

measure of the partitioning of a compound that is anticipated 

to occur between soil and water when the two phases are in 

contact. This parameter is quantified as the ratio between the 

concentrations of the compound in the soil (normalized to the 

organic carbon content of the soil) and in the aqueous solution 

at equilibrium. The Koc provides an indication of the extent 

of sorption of a compound to natural organic matter in the 

hydrologic system and, by extension, an inverse measure of 

the mobility of the compound in water within the subsurface.

Sorption The retention, through binding or association, of a 

solute ion or molecule by a solid. 

148 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Source water A stream, lake, other surface-water body, or 

aquifer from which water is drawn for human use.

Spatial extrapolation The use of statistical or other models 

to predict the value of a parameter (for example, the concentration of a chemical compound) in a location where it has not 

been measured.

Statistical model A model used to represent the effects of 

one or more processes of interest by quantitative, probabilistic 

relations (such as regressions) between one or more explanatory variables and a particular response variable.

Statistical significance The likelihood (commonly 

expressed as a probability, p) that the result of a statistical test 

may have occurred solely by chance. Observations associated 

with p values of 0.05 or less (a “95 percent or greater confidence level”) are typically deemed to be “statistically significant,” and thus, are unlikely to have occurred solely by chance.

Study Unit A major hydrologic system of the United States, 

geographically defined by surface- or ground-water features, 

in which NAWQA sampling studies are focused. The NAWQA 

studies during the first decade of assessments examined 51 

Study Units.

Subsurface The region of earth materials beneath the land 

surface that encompasses the soil, unsaturated, and saturated 

zones.

Subsurface tile-drain systems Perforated pipes that are buried in the ground to reduce the water content of poorly drained 

soils and divert shallow ground water to nearby streams.

Surface runoff The flow of water over the land surface, 

usually in response to intense rainfall or snowmelt events, 

irrigation, or rainfall on saturated soils, snow, or impervious 

surfaces (such as pavement).

T

Target organism An organism that an active ingredient of a 

pesticide is designed to control or kill.

Technical DDT A commercial DDT formulation that commonly contained approximately 80 percent p,p'-DDT, approximately 20 percent o,p'-DDT, and small amounts of o,o'-DDT, 

m,p'-DDT, p,p'-DDD, and other manufacturing by-products.

Technical mixture A formulation of a commercial chemical 

product that usually contains minor amounts of manufacturing by-products or other compounds in addition to the compound of interest (such as the active ingredient in a pesticide 

product).

Termiticide Pesticides that are used to kill termites, usually 

in buildings and other structures.

Time-weighted 95th percentile concentration The concentration of a given compound that is exceeded 5 percent of the 

time, or about 18 days per year (generally not consecutive). 

Tolerance level The maximum permissible concentration of 

a pesticide or pesticide degradate allowed in or on foods or 

animal feed, for the protection of human health. (Also referred 

to as a “maximum pesticide residue level.”)

Total chlordane concentration The sum of the concentrations of multiple compounds derived from commercial chlordane formulations (including components of the original product and degradates) that might be present in an environmental 

sample. For NAWQA bed-sediment and fish-tissue analyses, 

this consisted of the cis and trans isomers of both chlordane 

and nonachlor, as well as the chlordane degradate oxychlordane. For the chemical analyses of lake sediment cores by Van 

Metre and Mahler (2005), only cis-chlordane, trans-chlordane 

and trans-nonachlor were included.

Total DDT concentration The sum of the concentrations of 

all compounds of interest derived from DDT that might be 

present in an environmental sample. For NAWQA bed-sediment and fish-tissue analyses, this consisted of the p,p' and 

o,p' isomers of DDT, DDD, and DDE. For the chemical analyses of lake sediment cores by Van Metre and Mahler (2005), 

only the p,p' isomers of DDT, DDD, and DDE were included.

Total detection frequency The percentage of samples in 

which any of the analytes of interest are measured at any concentration without correcting to a common detection level.

Toxicity The degree to which the presence of a chemical substance at a particular concentration in a given environmental 

medium may be harmful to the health of humans and other 

organisms that come in contact with that medium.

Toxicity value A quantitative measure of the dose-response 

relationship observed in a test of the physiological effect of a 

particular chemical on a specific organism. Examples include 

LC50 and NOAEC values. 

Transformation The conversion of one compound to another 

through either abiotic or biotic processes.

Transformation product See degradate.

Triazine herbicides A group of pesticides—all sharing a 

six-membered aromatic ring consisting of three nitrogen atoms 

and three carbon atoms in an alternating sequence (a “symmetrical triazine ring”)—used primarily for weed control on 

corn, sorghum, cotton, sugarcane, orchards, fallow land, sod, 

rights-of-way, lawns, golf courses, and Christmas tree farms.

Tritium unit (TU) A measure of the concentration of tritium 

(

3

H), equal to 1 3H atom in 1018 atoms of hydrogen (H), or 3.24 

picocuries per liter (pCi/L).

U

Undeveloped stream A stream draining a watershed with 

25 percent or less of agricultural land and 5 percent or less of 

urban land.

Unique mixture A combination of two or more specific 

compounds detected in an environmental sample, regardless of 

whether other compounds are detected in the same sample.

Universal Soil Loss Equation An empirical equation developed by the U.S. Department of Agriculture to predict the 

Glossary 149
average amount of soil lost from a given location per year 

through wind and water erosion. The equation employs a 

variety of parameters related to rainfall, soil properties, and 

topography as explanatory variables.

Unsaturated zone The subsurface region of earth materials 

above the water table in which the pore spaces may contain a 

combination of air and water.

Urban stream A stream draining a watershed with more than 

5 percent of residential, commercial, transportation, urban 

recreational areas, and (or) industrial land, and 25 percent or 

less of agricultural land.

Use intensity The total quantity of a pesticide applied over 

a specified area, expressed in terms of the amount applied per 

unit area.

V

Volatilize To move spontaneously from a liquid or dissolved 

state to a gaseous state.

W

Water table The point below the land surface at which 

ground water is first encountered and below which the earth is 

saturated. 

Water-quality benchmark A threshold value above which the 

concentration of a specific chemical in a particular environmental medium may have adverse effects on human health, 

aquatic life, or fish-eating wildlife, and below which there is 

a low likelihood of such effects (see also aquatic-life benchmark, human-health benchmark, and wildlife benchmark).

Watershed The land area that drains into a particular stream, 

river, lake, estuary, or coastal zone.

Wildlife benchmark A threshold value above which the 

concentration of a chemical in water or fish tissue may have 

adverse effects on fish-eating wildlife.

150 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Appendixes 1–4

Appendix 1 153

Appendix 1—Pesticide compounds analyzed in NAWQA samples.

Table A. Pesticide compounds analyzed in NAWQA water samples.—Continued

[Pesticide compounds include pesticides, degradates, and manufacturing by-products. Pesticide compounds are grouped by pesticide class. Common synonyms 

are listed in parentheses in column 1. The cited references are listed by number at the end of Appendix 1. CAS, Chemical Abstracts Service; GCMS, gas chromatography/mass spectrometry; HPLC, high performance liquid chromatography; Parameter code, the number used to identify a pesticide in the U.S. Geological Survey National Water Information System and the U.S. Environmental Protection Agency Data Storage and Retrieval System; μg/L, micrograms per liter.]

Pesticide compound

 (synonym)

Type of pesticide compound

(parent pesticide, 

if degradate)

CAS number Parameter 

code

Analytical 

method

Long-term method 

detection level 1

(μg/L)

Amides

Acetochlor Herbicide 34256-82-1 49260 GCMS 0.003

Alachlor Herbicide 15972-60-8 46342 GCMS 0.002

2,6-Diethylaniline Degradate (Alachlor) 579-66-8 82660 GCMS 0.003

Metolachlor Herbicide 51218-45-2 39415 GCMS 0.006

Napropamide Herbicide 15299-99-7 82684 GCMS 0.003

Pronamide (Propyzamide) Herbicide 23950-58-5 82676 GCMS 0.002

Propachlor Herbicide 1918-16-7 04024 GCMS 0.005

Propanil Herbicide 709-98-8 82679 GCMS 0.005

Carbamates

Aldicarb Insecticide 116-06-3 49312 HPLC 0.100

Aldicarb sulfone Degradate (Aldicarb) 1646-88-4 49313 HPLC 0.100

Aldicarb sulfoxide Degradate (Aldicarb) 1646-87-3 49314 HPLC 0.140

Butylate Herbicide 2008-41-5 04028 GCMS 0.001

Carbaryl Insecticide 63-25-2 82680 GCMS 0.021

Carbofuran Insecticide 1563-66-2 82674 GCMS 0.010

EPTC Herbicide 759-94-4 82668 GCMS 0.001

3-Hydroxycarbofuran Degradate (Carbofuran) 16655-82-6 49308 HPLC 0.050

Methiocarb Insecticide 2032-65-7 38501 HPLC 0.030

Methomyl Insecticide 16752-77-5 49296 HPLC 0.240

Molinate Herbicide 2212-67-1 82671 GCMS 0.001

Oxamyl Insecticide 23135-22-0 38866 HPLC 0.080

Pebulate Herbicide 1114-71-2 82669 GCMS 0.002

Propham Herbicide 122-42-9 49236 HPLC 0.110

Propoxur (Baygon) Insecticide 114-26-1 38538 HPLC 0.060

Thiobencarb Herbicide 28249-77-6 82681 GCMS 0.002

Triallate Herbicide 2303-17-5 82678 GCMS 0.001

Chlorobenzoic acid esters

Dacthal (DCPA) 2 Herbicide 1861-32-1 82682 GCMS 0.002

Dacthal monoacid Degradate (Dacthal) 887-54-7 49304 HPLC 0.040

Chlorophenoxy acids

2,4-D Herbicide 94-75-7 39732 HPLC 0.080

2,4-DB Herbicide 94-82-6 38746 HPLC 0.130

Dichlorprop Herbicide 120-36-5 49302 HPLC 0.060
Table A. Pesticide compounds analyzed in NAWQA water samples.—Continued

[Pesticide compounds include pesticides, degradates, and manufacturing by-products. Pesticide compounds are grouped by pesticide class. Common synonyms 

are listed in parentheses in column 1. The cited references are listed by number at the end of Appendix 1. CAS, Chemical Abstracts Service; GCMS, gas chromatography/mass spectrometry; HPLC, high performance liquid chromatography; Parameter code, the number used to identify a pesticide in the U.S. Geological Survey National Water Information System and the U.S. Environmental Protection Agency Data Storage and Retrieval System; μg/L, micrograms per liter.]

Pesticide compound

 (synonym)

Type of pesticide compound

(parent pesticide, 

if degradate)

CAS number Parameter 

code

Analytical 

method

Long-term method 

detection level 1

(μg/L)

MCPA Herbicide 94-74-6 38482 HPLC 0.100

MCPB Herbicide 94-81-5 38487 HPLC 0.130

2,4,5-T Herbicide 93-76-5 39742 HPLC 0.040

2,4,5-TP (Silvex) Herbicide 93-72-1 39762 HPLC 0.030

Dinitroanilines

Benfluralin Herbicide 1861-40-1 82673 GCMS 0.005

Ethalfluralin Herbicide 55283-68-6 82663 GCMS 0.005

Oryzalin Herbicide 19044-88-3 49292 HPLC 0.140

Pendimethalin Herbicide 40487-42-1 82683 GCMS 0.011

Trifluralin Herbicide 1582-09-8 82661 GCMS 0.005

Miscellaneous

Bentazon Herbicide 25057-89-0 38711 HPLC 0.030

Norflurazon Herbicide 27314-13-2 49293 HPLC 0.021

Miscellaneous acids

Acifluorfen Herbicide 50594-66-6 49315 HPLC 0.040

Chloramben methyl ester Herbicide 7286-84-2 61188 HPLC 0.110

Clopyralid Herbicide 1702-17-6 49305 HPLC 0.210

Dicamba Herbicide 1918-00-9 38442 HPLC 0.050

Picloram Herbicide 1918-02-1 49291 HPLC 0.040

Nitrophenols

Dinoseb Herbicide 88-85-7 49301 HPLC 0.040

DNOC Herbicide 534-52-1 49299 HPLC 0.130

Organochlorines

Chlorothalonil Fungicide 1897-45-6 49306 HPLC 0.070

p,p'-DDE Degradate (p,p'-DDT) 72-55-9 34653 GCMS 0.001

Dichlobenil Herbicide 1194-65-6 49303 HPLC 0.050

Dieldrin Insecticide, Degradate 

(Aldrin)

60-57-1 39381 GCMS 0.002

alpha-HCH Degradate (gamma-HCH),

By-product in technical 

lindane 3

319-84-6 34253 GCMS 0.002

gamma-HCH (Lindane) Insecticide 58-89-9 39341 GCMS 0.002

Triclopyr Herbicide 55335-06-3 49235 HPLC 0.040

Organophosphates

Azinphos-methyl (Guthion) Insecticide 86-50-0 82686 GCMS 0.020

Chlorpyrifos Insecticide 2921-88-2 38933 GCMS 0.003

Diazinon Insecticide 333-41-5 39572 GCMS 0.003

154 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Table A. Pesticide compounds analyzed in NAWQA water samples.—Continued

[Pesticide compounds include pesticides, degradates, and manufacturing by-products. Pesticide compounds are grouped by pesticide class. Common synonyms 

are listed in parentheses in column 1. The cited references are listed by number at the end of Appendix 1. CAS, Chemical Abstracts Service; GCMS, gas chromatography/mass spectrometry; HPLC, high performance liquid chromatography; Parameter code, the number used to identify a pesticide in the U.S. Geological Survey National Water Information System and the U.S. Environmental Protection Agency Data Storage and Retrieval System; μg/L, micrograms per liter.]

Pesticide compound

 (synonym)

Type of pesticide compound

(parent pesticide, 

if degradate)

CAS number Parameter 

code

Analytical 

method

Long-term method 

detection level 1

(μg/L)

Disulfoton Insecticide 298-04-4 82677 GCMS 0.011

Ethoprop (Ethoprophos) Insecticide 13194-48-4 82672 GCMS 0.002

Fonofos Insecticide 944-22-9 04095 GCMS 0.001

Malathion Insecticide 121-75-5 39532 GCMS 0.014

Parathion (Ethyl parathion) Insecticide 56-38-2 39542 GCMS 0.005

Parathion-methyl (Methyl parathion) Insecticide 298-00-0 82667 GCMS 0.003

Phorate Insecticide 298-02-2 82664 GCMS 0.006

Terbufos Insecticide 13071-79-9 82675 GCMS 0.009

Phenols

Bromoxynil Herbicide 1689-84-5 49311 HPLC 0.030

 Pyrethroids

cis-Permethrin 2 Insecticide 54774-45-7 82687 GCMS 0.003

 Sulfite esters

Propargite Acaricide 2312-35-8 82685 GCMS 0.011

Triazines

Atrazine Herbicide 1912-24-9 39632 GCMS 0.004

Cyanazine Herbicide 21725-46-2 04041 GCMS 0.009

Deethylatrazine Degradate (Atrazine) 6190-65-4 04040 GCMS 0.003

Metribuzin Herbicide 21087-64-9 82630 GCMS 0.003

Prometon Herbicide 1610-18-0 04037 GCMS 0.007

Simazine Herbicide 122-34-9 04035 GCMS 0.006

Uracils

Bromacil Herbicide 314-40-9 04029 HPLC 0.040

Terbacil Herbicide 5902-51-2 82665 GCMS 0.017

Ureas

Diuron Herbicide 330-54-1 49300 HPLC 0.060

Fenuron Herbicide 101-42-8 49297 HPLC 0.030

Fluometuron Herbicide 2164-17-2 38811 HPLC 0.030

Linuron Herbicide 330-55-2 82666 GCMS 0.018

Neburon Herbicide 555-37-3 49294 HPLC 0.030

Tebuthiuron Herbicide 34014-18-1 82670 GCMS 0.008

1

 The long-term method detection level (reference 1) is calculated annually. The value reported in the table is the maximum long-term method detection level 

for the period 1992–2001.

2

 This pesticide also can be considered an organochlorine pesticide because it is an organic pesticide with multiple chlorine substituents.

3

 Prior to 1977, alpha-HCH was a manufacturing by-product in technical lindane, which is a mixture of several isomers of hexachlorocyclohexane 

(reference 2). 

Appendix 1 155
Table B. Pesticide compounds analyzed in NAWQA bed-sediment or whole-fish samples.—Continued

[Pesticide compounds include pesticides, degradates, and manufacturing by-products. Pesticide compounds are grouped by pesticide class. Common synonyms 

are listed in parentheses in column 1. The cited references are listed by number at the end of Appendix 1. CAS, Chemical Abstracts Service; Parameter code, 

the number used to identify a pesticide in the U.S. Geological Survey National Water Information System and the U.S. Environmental Protection Agency Data 

Storage and Retrieval System; μg/kg dw, micrograms per kilogram dry weight; μg/kg ww, micrograms per kilogram wet weight; —, not analyzed.]

Pesticide compound 

(synonym)

Type of pesticide compound

 (parent pesticide, if degradate)

CAS 

number

Bed sediment Whole fish

Parameter 

code

Reporting 

level 

(μg/kg dw)

Parameter 

code

Reporting 

level 

(μg/kg ww)

Chlorobenzoic acid esters

Dacthal (DCPA) 1 Herbicide 1861-32-1 49324 5 49378 5

Organochlorines

Aldrin Insecticide, Component of total dieldrin 2 309-00-2 49319 1 49353 5

cis-Chlordane Insecticide, Component of total chlordane 3 5103-71-9 49320 1 49380 5

trans-Chlordane Insecticide, Component of total chlordane 3 5103-74-2 49321 1 49379 5

Chloroneb Insecticide 2675-77-6 49322 5 — —

o,p′-DDD Degradate (o,p′-DDT), Component of total 

DDT 4

53-19-0 49325 1 49374 5

p,p′-DDD (p,p′-TDE) Insecticide, Degradate (p,p′-DDT), Component of total DDT 4

72-45-8 49326 1 49375 5

o,p′-DDE Degradate (o,p′-DDT), Component of total 

DDT 4

3424-82-6 49327 1 49373 5

p,p′-DDE Degradate (p,p′-DDT), Component of total 

DDT 4

72-55-9 49328 1 49372 5

o,p′-DDT By-product in technical DDT, Component of 

total DDT 4

789-02-6 49329 2 49377 5

p,p′-DDT Insecticide, Component of total DDT 4 50-29-3 49330 2 49376 5

Dieldrin Insecticide, Degradate (Aldrin), Component 

of total dieldrin 2

60-57-1 49331 1 49371 5

Endosulfan I (alphaEndosulfan)

Insecticide 959-98-8 49332 1 — —

Endrin Insecticide 72-20-8 49335 1 49370 5

alpha-HCH Degradate (gamma-HCH), By-product in 

technical lindane, Component of total 

HCH 5

319-84-6 49338 1 49366 5

beta-HCH By-product in technical lindane, Component 

of total HCH 5

319-85-7 49339 1 49365 5

gamma-HCH (Lindane) Insecticide, Component of total HCH 5 58-89-9 49345 1 49363 5

delta-HCH Degradate (gamma-HCH), By-product in 

technical lindane, Component of total 

HCH 5

319-86-8 — — 49364 5

Heptachlor Insecticide, Component of total heptachlor 6 76-44-8 49341 1 49369 5

Heptachlor epoxide Degradate (Heptachlor), Component of total 

heptachlor 6

1024-57-3 49342 1 49368 5

Hexachlorobenzene Insecticide 118-74-1 49343 1 49367 5

Isodrin Insecticide 465-73-6 49344 1 — —

o,p′-Methoxychlor Insecticide, Component of total methoxychlor 7

30667-99-3 49347 5 49362 5

p,p′-Methoxychlor Insecticide, Component of total methoxychlor 7

72-43-5 49346 5 49361 5

Mirex Insecticide 2385-85-5 49348 1 49360 5

cis-Nonachlor By-product in technical chlordane, Component of total chlordane 3

5103-73-1 49316 1 49359 5

156 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Table B. Pesticide compounds analyzed in NAWQA bed-sediment or whole-fish samples.—Continued

[Pesticide compounds include pesticides, degradates, and manufacturing by-products. Pesticide compounds are grouped by pesticide class. Common synonyms 

are listed in parentheses in column 1. The cited references are listed by number at the end of Appendix 1. CAS, Chemical Abstracts Service; Parameter code, 

the number used to identify a pesticide in the U.S. Geological Survey National Water Information System and the U.S. Environmental Protection Agency Data 

Storage and Retrieval System; μg/kg dw, micrograms per kilogram dry weight; μg/kg ww, micrograms per kilogram wet weight; —, not analyzed.]

Pesticide compound 

(synonym)

Type of pesticide compound

 (parent pesticide, if degradate)

CAS 

number

Bed sediment Whole fish

Parameter 

code

Reporting 

level 

(μg/kg dw)

Parameter 

code

Reporting 

level 

(μg/kg ww)

trans-Nonachlor By-product in technical chlordane, Component of total chlordane 3

39765-80-5 49317 1 49358 5

Oxychlordane Degradate (Chlordane), Component of total 

chlordane 3

27304-13-8 49318 1 49357 5

Pentachloroanisole Degradate (Pentachlorophenol) 1825-21-4 49460 1 49356 5

Toxaphene Insecticide, Technical mixture 8001-35-2 49351 200 49355 200

Pyrethroids

cis-Permethrin 1 Insecticide, Component of total permethrin 8 52774-45-7 49349 5 — —

trans-Permethrin 1 Insecticide, Component of total permethrin 8 51877-74-8 49350 5 — —

1

 This pesticide also can be considered an organochlorine pesticide because it is an organic pesticide with multiple chlorine substituents.

2

 The pesticide group “total dieldrin” refers to the summed concentrations of aldrin and dieldrin.

3

 The pesticide group “total chlordane” refers to the summed concentrations of cis-chlordane, trans-chlordane, cis-nonachlor, trans-nonachlor, and oxychlordane. Chlordane was applied as a technical-grade mixture of over 140 compounds, including nonachlor isomers and other manufacturing by-products.

4

 The pesticide group “total DDT” refers to the summed concentrations of o,p′-DDT, p,p′-DDT, o,p′-DDD, p,p′-DDD, o,p′-DDE, and p,p′-DDE. Technical 

DDT contained p,p′-DDT (the active ingredient) and o,p′-DDT (a manufacturing by-product).

5

 The pesticide group “total HCH” refers to the summed concentrations of alpha-HCH, beta-HCH, gamma-HCH, and delta-HCH. Technical lindane is a mixture of several isomers of hexachlorocyclohexane; gamma-HCH is the active ingredient, and the other isomers are manufacturing by-products. In 1977, USEPA 

cancelled inclusion of alpha- and beta-HCH in technical-grade lindane (reference 2). 

6

 The pesticide group “total heptachlor” refers to the summed concentrations of heptachlor and heptachlor epoxide.

7

 The pesticide group “total methoxychlor” refers to the summed concentrations of o,p′-methoxychlor and p,p′-methoxychlor.

8

 The pesticide group “total permethrin” refers to the summed concentrations of cis-permethrin and trans-permethrin.

References—Appendix 1

1. Oblinger Childress, C.J., Foreman, W.T., Conner, B.F., and 

Maloney, T.J., 1999, New reporting procedures based on 

long-term method detection levels and some considerations 

of interpretations of water-quality data provided by the U.S. 

Geological Survey National Water Quality Laboratory: U.S. 

Geological Survey Open-File Report 99-193, 19 p.

2. U.S. Environmental Protection Agency, 1992, National 

study of chemical residues in fish: U.S. Environmental 

Protection Agency, Office of Science and Technology, EPA823-R-92-008b, v. 2, variously paged.

Appendix 1 157

Appendix 2 159

Appendix 2—Properties affecting transport and fate.

Table A. Properties affecting the transport and fate of selected pesticide compounds.—Continued

[Pesticide compounds selected are those detected most frequently in NAWQA samples (see figs. 4–2 and 4–4), as well as several that were detected infrequently, 

despite extensive use. All values meas­ured at (or estima­ted for) 25°C, ex­cept for those shown in italics. Unless noted otherwise, (1) values for octanol-water partition coefficient (Kow, dimen­sion­less), soil organic carbon-water partition coefficient (Koc), water solubility (Sw) and Henry’s law constant (KH) are from Mackay 

and others (1997); (2) transfor­ma­tion half-lives in soil and water were measured in the laboratory (rather than in the field) at neutral pH in the dark, and obtained 

from the U.S. Department of Agriculture (2005); and (3) all are re­com­men­ded values selec­ted by the com­pi­lation authors when more than one value was available from the litera­ture. Com­­­pounds are listed in the same order as in figures 4–2 and 4–4. Numbers of signi­ficant figures are iden­tical to those given in original 

sources. mg/L, milligrams per liter; mL/g, milliliters per gram; NA, data not available from any of the references consulted; Pa•m3/mol, pascal-cubic meters per 

mole; >, greater than.]

Pesticide compound 

(synonym) log Kow

log Koc

(Koc in mL/g)

Sw

(mg/L)

log KH

(KH in Pa•m3/

mol)

Half-life for transformation

(days)

In aerobic soil In water

Agricultural herbicides and degradates detected most frequently in water

Atrazine 2.75 2.00 30 -3.54 146 1742

Deethylatrazine 1,21.3 1,21.90 12,700 1,2-4.12 1,2170 NA

Metolachlor 3.13 2.26 430 -2.63 26 1,2410

Cyanazine 2.22 2.3 171 -6.52 2

17 2>200

Alachlor 2.8 2.23 240 -2.7 1,220.4 1,2640

Acetochlor 13.0 1,22.38 1,3223 1,3-2.15 1,2,314 12,300

Metribuzin 1,31.60 1.72 1,31,000 1,2-5.31 172 2>200

Bentazon 1,22.80 1,21.54 1,3500 1,2,3-3.7 1,235 1,2,3>200

EPTC 3.2 2.3 370 0.00988 1,37 2>200

Trifluralin 5.34 4.14 20.5 1.00 169 2>32

Molinate 3.21 1.92 970 -0.839 1,321 2>200

Norflurazon1 2.45 2.55 34 -4.46 130 2>200

Urban herbicides detected most frequently in water

Simazine 2.18 2.11 5 -3.46 291 1,2>32

Prometon 2.99 2.54 750 -4.05 932 2>200

Tebuthiuron 1,21.79 12.1 1,22,400 1-4.88 11,050 2>2,700

2,4-D 2.81 11.68 890 -3.61 1,22.3 1,2732

Diuron 2.78 2.6 40 -3.17 372 >500

Dacthal (DCPA) 14.28 13.75 10.5 1-0.66 116 2>200

Bromacil 2.11 1.86 815 -4.89 275 2>30

Insecticides detected most frequently in water

Diazinon 3.3 2.76 60 -1.39 39 140

Chlorpyrifos 4.92 3.78 0.73 0.0374 30.5 29

Carbofuran 2.32 2.02 351 -4.30 11 2289

Carbaryl 2.36 2.36 120 -4.35 17 11

Malathion 2.8 3.26 145 -2.64 <1 26.3

Dieldrin 5.20 4.08 0.17 0.0492 NA 3,830

Organochlorine pesticide compounds detected most frequently in bed sediment and fish tissue

p,p'-DDE 5.7 5.0 0.04 0.900 NA 1,2,3>44,000

p,p'-DDD 5.5 5.0 0.05 -0.194 NA 2

10,000

p,p'-DDT 6.19 5.4 0.0055 0.37 NA 1,2,35,000
Table A. Properties affecting the transport and fate of selected pesticide compounds.—Continued

[Pesticide compounds selected are those detected most frequently in NAWQA samples (see figs. 4–2 and 4–4), as well as several that were detected infrequently, 

despite extensive use. All values meas­ured at (or estima­ted for) 25°C, ex­cept for those shown in italics. Unless noted otherwise, (1) values for octanol-water partition coefficient (Kow, dimen­sion­less), soil organic carbon-water partition coefficient (Koc), water solubility (Sw) and Henry’s law constant (KH) are from Mackay 

and others (1997); (2) transfor­ma­tion half-lives in soil and water were measured in the laboratory (rather than in the field) at neutral pH in the dark, and obtained 

from the U.S. Department of Agriculture (2005); and (3) all are re­com­men­ded values selec­ted by the com­pi­lation authors when more than one value was available from the litera­ture. Com­­­pounds are listed in the same order as in figures 4–2 and 4–4. Numbers of signi­ficant figures are iden­tical to those given in original 

sources. mg/L, milligrams per liter; mL/g, milliliters per gram; NA, data not available from any of the references consulted; Pa•m3/mol, pascal-cubic meters per 

mole; >, greater than.]

Pesticide compound 

(synonym) log Kow

log Koc

(Koc in mL/g)

Sw

(mg/L)

log KH

(KH in Pa•m3/

mol)

Half-life for transformation

(days)

In aerobic soil In water

o,p'-DDE 5.8 1,25.58 0.1 0.405 NA NA

o,p'-DDD 6.0 1,25.36 2

0.10 1-2.7 NA 2NA

o,p'-DDT 2

NA 2NA 0.026 -0.460 NA NA

cis-Chlordane 6.0 5.5 0.056 -0.466 NA 1

>7.2×107

trans-Chlordane 6.0 5.5 0.056 -0.582 NA 1,2>10,000

Nonachlor 1 5.66 4.86 0.06 -1.69 NA NA

Oxychlordane1 2.6 2.48 200 -1.52 NA NA

Dieldrin 5.20 4.08 0.17 0.0492 NA 3,830

Heptachlor epoxide 5.0 4.0 0.35 10.51 NA NA

Pentachloroanisole 1 5.66 4.62 0.2 2.91 NA NA

Hexachlorobenzene 15.31 14.7 10.0062 1,21.69 NA 1,2>26,000

Heavily used pesticides not detected frequently in water

Chlorothalonil 2.64 3.2 0.6 1.77 NA 2>200

Dicamba 2.21 11.11 4500 -3.66 1,228 2>200

Parathion-methyl (Methyl parathion) 3.0 3.7 25 -1.68 1,23.3 41

Pendimethalin 15.2 14.13 10.275 10.0899 1300 2>200

Terbufos 4.48 2.70 5 0.39 5 1.9

1

Value(s) obtained from sources other than Mackay and others (1997) for Kow, Koc, Sw, and KH; or U.S. Department of Agriculture (2005) for transformation 

half-lives. See http://ca.water.usgs.gov/pnsp/pubs/circ1291/ for data sources.

2

See http://ca.water.usgs.gov/pnsp/pubs/circ1291/ for details related to computation or selection of parameter value.

3

See http://ca.water.usgs.gov/pnsp/pubs/circ1291/ for temperature of measurement.

References—Appendix 2

Mackay, D., Shiu, W-Y., and Ma, K-C, 1997, Illustrated handbook of physical-chemical properties and environmental 

fate for organic chemicals, Volume V. Pesticide Chemicals: 

New York, NY, Lewis Publishers, 812 p.

U.S. Department of Agriculture, 2005, Agricultural Research 

Service (ARS), Pesticide Properties Database, accessed 

August 22, 2005, at http://www.ars.usda.gov/Services/docs.

htm?docid=6433.

160 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Appendix 3 161

Appendix 3— Water-quality benchmarks for pesticide compounds that exceeded one or more benchmarks in NAWQA samples.

Table A. Water-quality benchmarks for pesticide compounds that exceeded one or more benchmarks in NAWQA water samples.—Continued[For pesticide compounds in water, benchmarks are for protection of human health and aquatic life. This table includes only those pesticide compounds that exceeded one or more benchmarks in NAWQA water samples; for these compounds, all benchmark values that were compared with concentrations in water are listed here. The complete set of benchmarks used for all pesticide compounds analyzed by NAWQA is available at http://ca.water.usgs.gov/pnsp/pubs/circ1291/. Common synonyms are listed in parentheses in column 1. OPP references are listed by number at the end of Appendix 3. Environmental concentration, the measured or calculated concentration statistic that is appropriate for comparison with the benchmark; CRC(10-6), 10-6 cancer risk concentration; EC50, 50 percent effect concentration; EFED, Environmental Fate and Effects Division; USEPA, U.S. Environmental Protection Agency; HA-L, lifetime health advisory; IRED, Interim Reregistration Eligibility Decision; IRIS, Integrated Risk Information System database; LC50, 50 percent lethal concentration; LOC, level of concern; MCL, maximum contaminant level; NOAEC, no-observed-adverse-effects concentration; OPP, Office of Pesticide Programs; RED, Reregistration Eligibility Decision; μg/L, microgram sper liter; >, greater than; <, less than; —, no benchmark available.]

Pesticide compound 

(synonym)

Human-health benchmarks 

(μg/L) Aquatic-life benchmarks (μg/L)

Derived from USEPA drinking-water standards and 

guidelines (Office of Water)

Derived from USEPA 

ambient water-quality 

criteria for aquatic life

(Office of Water)

Derived from USEPA Reregistration Eligibility Decisions 

and ecological risk assessments 

(Office of Pesticide Programs) 

Value 1 Type Acute 2 Chronic 2 Acute fish 3 Chronic fish 4 Acute invertebrates 5 Chronic invertebrates 6 Acute nonvascular plants 7 Acute vascular plants 8 Aquatic community effects 9 OPP refer- ences

Environmental 

concentration

(μg/L): 

Timeweighted 

annual mean

Each 

individual 

sample

4-day 

moving 

average

Each 

individual 

sample

60-day 

moving 

average

Each 

individual 

sample

21-day 

moving 

average

Each 

individual 

sample

Each 

individual 

sample

60-day 

moving 

average

 Amides

Alachlor 2 MCL — — 900 187 1,600 110 1.64 — — (3)

 Carbamates

Carbaryl 700 HA-L — — 10 125 10 210 2.55 1.5 1,100 — — (4, 5)Carbofuran 40 MCL — — 44 5.7 1.115 0.75 — — — (6)Methomyl 200 HA-L — — 265 57 4.4 11 0.4 — — — (7)Molinate — — — — 105 12 210 170 12 340 220 3,300 — (8)

Thiobencarb — — — — 280 — 50 1.0 17 770 — (9)

 Nitrophenols

Dinoseb 7 MCL — — — — — — — — — —

 Organochlorines

p,p'-DDE 13 0.1 CRC(10-6) 14 1.1 14 0.001 — — — — — — — —Dieldrin 0.002 CRC(10-6) 0.24 0.056 — — — — — — — —gamma-HCH 

(Lindane)

0.2 MCL 0.95 — 0.85 12 1.7 0.5 12 1 — — — (11)
Table A. Water-quality benchmarks for pesticide compounds that exceeded one or more benchmarks in NAWQA water samples.—Continued

[For pesticide compounds in water, benchmarks are for protection of human health and aquatic life. This table includes only those pesticide compounds that exceeded one or more benchmarks in NAWQA water samples; for these compounds, all benchmark values that were compared with concentrations in water are listed here. The complete set of benchmarks used for all pesticide compounds analyzed by NAWQA is available at http://ca.water.usgs.gov/pnsp/pubs/circ1291/. Common synonyms are listed in parentheses in column 1. OPP references are listed by number at the end of Appendix 3. Environmental concentration, the measured or calculated concentration statistic that is appropriate for comparison with the benchmark; CRC(10-6), 10-6 cancer risk concentration; EC50, 50 percent effect concentration; EFED, Environmental Fate and Effects Division; USEPA, U.S. Environmental Protection Agency; HA-L, lifetime health advisory; IRED, Interim Reregistration Eligibility Decision; IRIS, Integrated Risk Information System database; LC50, 50 percent lethal concentration; LOC, level of concern; MCL, maximum contaminant level; NOAEC, no-observed-adverse-effects concentration; OPP, Office of Pesticide Programs; RED, Reregistration Eligibility Decision; μg/L, microgram sper liter; >, greater than; <, less than; —, no benchmark available.]

Pesticide compound 

(synonym)

Human-health benchmarks 

(μg/L) Aquatic-life benchmarks (μg/L)

Derived from USEPA drinking-water standards and 

guidelines (Office of Water)

Derived from USEPA 

ambient water-quality 

criteria for aquatic life

(Office of Water)

Derived from USEPA Reregistration Eligibility Decisions 

and ecological risk assessments 

(Office of Pesticide Programs) 

Value 1 Type Acute 2 Chronic 2 Acute fish 3 Chronic fish 4 Acute invertebrates 5 Chronic invertebrates 6 Acute nonvascular plants 7 Acute vascular plants 8 Aquatic community effects 9 OPP refer- ences

Environmental 

concentration

(μg/L): 

Timeweighted 

annual mean

Each 

individual 

sample

4-day 

moving 

average

Each 

individual 

sample

60-day 

moving 

average

Each 

individual 

sample

21-day 

moving 

average

Each 

individual 

sample

Each 

individual 

sample

60-day 

moving 

average

 Organophosphates

Azinphos-methyl 

(Guthion)

— — — 0.01 0.18 12 0.36 0.08 12 0.16 — — — (12)

Chlorpyrifos 20 HA-L 0.083 0.041 0.9 0.57 0.05 0.04 140 — — (13, 14)Diazinon 0.6 HA-L — — 45 15 0.55 10 0.1 10 0.17 3,700 — — (15, 16)Disulfoton 0.3 HA-L — — 19.5 12 39 1.95 0.037 — — — (17)Malathion 100 HA-L — 0.1 2 12 4 0.25 0.06 — — — (18)Parathion (Ethyl 

parathion)

— — 0.065 0.013 9 16 0.19 0.02 0.002 — — — (19)

Parathion-methyl 

(Methyl parathion)

2 HA-L — — 500 15 80 0.07 0.02 5,300 — — (20)

Phorate — — — — 0.5 12 1 0.30 0.21 1,300 — — (21, 22)Terbufos 0.9 HA-L — — 0.385 12 0.77 0.1 0.030 — — — (23)

 Sulfite esters

Propargite — — — — 10 15.5 10 16 37 9 19.4 75,000 — (24)

 Triazines

Atrazine 3 MCL — — 2,650 62 360 62 32 18 17.5 (25, 26)Cyanazine 1 HA-L — — — — — — — — — —

 Ureas

Diuron 10 HA-L — — 355 26 80 12 160 2.4 — — (27)

162 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
1 From reference 1, unless noted otherwise.

2 From reference 2.3 Benchmark = Toxicity value x LOC. For acute fish, the toxicity value is generally the lowest 96-hour LC50 in a standardized test (usually with rainbow trout, fathead minnow, or bluegill), and the LOC is 0.5.4 Benchmark = Toxicity value x LOC. For chronic fish, the toxicity value is usually the lowest NOEAC from a life-cycle or early life-stage test (usually with rainbow trout or fathead minnow), and the LOC is 1.5 Benchmark = Toxicity value x LOC. For acute invertebrate, the toxicity value is usually the lowest 48- or 96-hour EC50 or LC50 in a standardized test (usually with midge, scud, or daphnids), and the LOC is 0.5.6 Benchmark = Toxicity value x LOC. For chronic invertebrates, the toxicity value is usually the lowest NOAEC from a life-cycle test with invertebrates (usually with midge, scud, or daphnids), and the LOC is 1.7 Benchmark = Toxicity value x LOC. For acute nonvascular plants, the toxicity value is usually a short-term (less than 10 days) EC50 (usually with green algae or diatoms), and the LOC is 1.8 Benchmark = Toxicity value x LOC. For acute vascular plants, the toxicity value is usually a short-term (less than 10 days) EC50 (usually with duckweed) and the LOC is 1.9 Exceedance of this benchmark concentration, as an average for any 60-day period, could cause community-level effects on aquatic plants (based on changes in plant community diversity as predicted by the 

Comprehensive Aquatic Systems Model), and indirect effects on fish and aquatic invertebrates from disturbance of the aquatic plant community (from reference 26).

10 Although the underlying acute toxicity value is greater than the chronic toxicity value, the acute benchmark is lower than the chronic benchmark because acute and chronic toxicity values were multiplied by 

LOC values of 0.5 and 1, respectively.

11 Because the underlying toxicity value is a “greater-than” value (such as >265,000), this benchmark may overestimate toxicity.12 The chronic benchmark is based on the acute toxicity value (which was lower than the lowest available chronic toxicity value) and, therefore, may underestimate chronic toxicity.13 From reference 10.14 Benchmark applies to total DDT, so comparison with measured p,p'-DDE concentration may underestimate potential effects. To account for potential contamination of some stream-water samples by p,p'-DDE, 

which may have resulted from sample collection and analysis procedures, exceedance of the chronic aquatic-life criterion was determined after subtracting 0.002 ug/L from the measured concentration of p,p'-DDE. The expected overall effect of this compensation is to underestimate the true frequency of exceedance of the 0.001 ug/L chronic aquatic-life criterion for p,p'-DDE (Details are provided at http://ca.water.usgs.gov/pnsp/pubs/circ1291/.)

15 Because the underlying toxicity value is a “less-than” value (such as <1,500), this benchmark may underestimate toxicity.16 This benchmark has greater uncertainty than usual because of methods used or conditions in the underlying toxicity study.

Appendix 3 163
Table B. Water-quality benchmarks for pesticide compounds that exceeded one or more benchmarks in NAWQA bed sediment or whole-fish samples.

[For pesticide compounds in bed sediment and whole-fish tissue, benchmarks are for protection of aquatic life and fish-eating wildlife, respectively. This table includes only those pesticide compounds that exceeded one or more benchmarks in NAWQA sediment or fish samples; for these compounds, all benchmark values that were compared with concentrations in sediment or fish are listed here. The complete set of benchmarks used for all pesticide compounds analyzed by NAWQA is available at http://ca.water.usgs.gov/pnsp/pubs/circ1291/. Common synomyms are listed in parentheses in column 1. The cited references are listed by number at the end of Appendix 3. BenchmarkHigh, highest value in the range of wildlife benchmarks available for a given pesticide compound or group; BenchmarkLow, lowest value in the range of wildlife benchmarks available for a given pesticide or group; Cf, concentration in food; Eisler-PC, proposed criterion from Contaminant Hazard Review series by R. Eisler and colleagues; ESB, equilibrium partitioning sediment benchmark; FFC, fish flesh criterion; NOAEL-ECF, no-observed-adverse-effects level equivalent concentration in food; TEC, threshold effect concentration; TRG, tissue residue guideline; USEPA, U.S. Environmental Protection Agency; μg/g, micrograms per gram; μg/kg dw, micrograms per kilogram dry weight; μg/kg ww, micrograms per kilogram wet weight; —, no benchmark available.]

Pesticide compound 

(synonym)

Bed-sediment benchmarks for protection of benthic 

aquatic organisms Whole-fish benchmarks for protection of fish-eating wildlife

Consensus-based threshold effect concentration 

(TEC) (μg/kg dw) 1

USEPA equilibrium 

partitioning sediment 

benchmark (ESB) (μg/g of 

sediment organic carbon) 1

BenchmarkLow value 

(μg/kg ww) 2 Type of BenchmarkLow BenchmarkHigh value (μg/kg ww) 2 Type of BenchmarkHigh

Organochlorines

Total Chlordane 3 3.24 — 300 Eisler-PC 4,200 NOAEL-ECFo,p'-DDD + p,p'-DDD 4 4.88 — see (5) — see (5) —o,p'-DDE + p,p'-DDE 4 3.16 — see (5) — see (5) —

o,p'-DDT + p,p'-DDT 4 4.16 — see (5) — see (5) —Total DDT 5 5.28 — 6 NOAEL-ECF 200 New York FFC

Dieldrin 1.90 12 81 NOAEL-ECF 6 120 New York FFC

Endosulfan I (alphaEndosulfan — 0.29 — — — —

Endrin 2.22 5.4 20 NOAEL-ECF 25 New York FFCgamma-HCH (Lindane) 2.37 0.37 3,950 NOAEL-ECF 3,950 NOAEL-ECF

Heptachlor epoxide 2.47 — see (7) — see (7) —Total Heptachlor 7 — — 200 New York FFC 529 NOAEL-ECF

Total Methoxychlor 8 — 1.9 16,300 NOAEL-ECF 16,300 NOAEL-ECFToxaphene — 10 6.3 Canadian TRG 32,500 NOAEL-ECF

1 TECs are from reference 28. ESBs are from references 29–31.2 BenchmarkLow and BenchmarkHigh refer to the range of benchmark values from the following four sources: Oak Ridge NOAEL-equivalent concentrations in food (the lowest Cf value for piscivorous wildlife 

species from reference 32); New York fish flesh criteria—noncancer (reference 33); Canadian tissue residue guidelines (references 34–35); proposed criteria from the U.S. Fish and Wildlife Service’s report series, Contaminant Hazard Reviews (references 36–37).

3 Sum of the concentrations of cis- and trans-chlordane, cis- and trans-nonachlor, and oxychlordane.4 Benchmark applies to the sum of these two compounds.5 Sum of the concentrations of o,p' and p,p' isomers of DDD, DDE, and DDT.6 The benchmark is for total dieldrin (sum of the concentrations of aldrin plus dieldrin). However, this benchmark is considered applicable to dieldrin concentration data because only one fish sample contained any aldrin (10 μg/kg), and in that sample both aldrin and dieldrin (43 μg/kg) were well below all benchmarks.7 Sum of the concentrations of heptachlor and heptachlor epoxide.8 Sum of the concentrations of o,p' and p,p' isomers of methoxychlor.

164 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
References—Appendix 3

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2004 Edition of the drinking water standards and health 

advisories: U.S. Environmental Protection Agency, Office 

of Water, EPA-822-R-04-005, accessed June 9, 2005, 

at http://www.epa.gov/waterscience/criteria/drinking/

standards/dwstandards.pdf.

2. USEPA (U.S. Environmental Protection Agency), 2004, 

National recommended water quality criteria: U.S. Environmental Protection Agency, Office of Water, accessed 

June 9, 2005, at http://www.epa.gov/waterscience/criteria/

nrwqc-2004.pdf.

3. USEPA (U.S. Environmental Protection Agency), 1998, 

Reregistration eligibility decision (RED) alachlor: U.S. 

Environmental Protection Agency, Prevention, Pesticides 

and Toxic Substances, EPA 738-R-98-020, December 

1998, accessed September 14, 2005, at http://www.epa.

gov/oppsrrd1/REDs/0063.pdf.

4. USEPA (U.S. Environmental Protection Agency), 2003, 

Environmental fate and ecological risk assessment for the 

reregistration of carbaryl: U.S. Environmental Protection 

Agency, Prevention, Pesticides and Toxic Substances, 

Public Docket Number EPA-HQ-OPP-2003-010-0005, 

accessed February 1, 2006, at http://www.regulations.gov.

5. USEPA (U.S. Environmental Protection Agency), 2004, 

Interim reregistration eligibility decision for carbaryl 

(revised October 22, 2004), list A, case 0080: U.S. Environmental Protection Agency, Office of Pesticide Programs, accessed September 22, 2005, at http://www.epa.

gov/oppsrrd1/REDs/carbaryl_ired.pdf.

6. USEPA (U.S. Environmental Protection Agency), 2005, 

Reregistration eligibility science chapter for carbofuran, 

environmental fate and effects chapter, June 3, 2005, 

Appendix 1, Carbofuran deterministic environmental risk 

assessment: U.S. Environmental Protection Agency, Public 

Docket Number EPA-HQ-OPP-2005-0162-0006, accessed 

February 1, 2006, at http://www.regulations.gov.

7. USEPA (U.S. Environmental Protection Agency), 1998, 

Reregistration eligibility decision (RED) methomyl: U.S. 

Environmental Protection Agency, Prevention, Pesticides 

and Toxic Substances, EPA 738-R-98-021 December 

1998, accessed October 11, 2005, at http://www.epa.gov/

oppsrrd1/REDs/0028red.pdf.

8. USEPA (U.S. Environmental Protection Agency), 2001a, 

Revised EFED reregistration eligibility chapter for 

molinate, Appendix C, ecological toxicity assessment: 

U.S. Environmental Protection Agency, accessed September 16, 2005, at http:// www.epa.gov/oppsrrd1/reregistration/molinate/AppCecologicaltoxicityasst.pdf.

9. USEPA (U.S. Environmental Protection Agency), 1997, 

Reregistration eligibility decision (RED) thiobencarb: U.S. 

Environmental Protection Agency, Prevention, Pesticides 

and Toxic Substances, EPA 738-R-97-013 December 

1997, accessed October 5, 2005, at http://www.epa.gov/

oppsrrd1/REDs/2665red.pdf.

10. USEPA (U.S. Environmental Protection Agency), 2005, 

Integrated Risk Information System (IRIS) database: U.S. 

Environmental Protection Agency, Office of Research and 

Development, National Center for Environmental Assessment, accessed June 9, 2005, at http://www.epa.gov/iris.

11. USEPA (U.S. Environmental Protection Agency), 2001, 

Lindane RED chapter: environmental fate and ecological 

risk assessment: seed treatment, Appendix I, Ecological 

effects data revised December 20, 2001: U.S. Environmental Protection Agency, Office of Pesticide Programs, Environmental Fate and Effects Division, accessed October 

10, 2005, at http://www.epa.gov/oppsrrd1/reregistration/

lindane/efed_ra_revised.pdf. 

12. USEPA (U.S. Environmental Protection Agency), 2005, 

Azinphos methyl insecticide: ecological risk assessment 

for the use of azinphos methyl on caneberries, cranberries, 

peaches, potatoes, and southern pine seeds (group 2 uses): 

U.S. Environmental Protection Agency, Office of Prevention, Pesticides and Toxic Substances, Public Docket 

Number EPA-HQ-OPP-2005-0061-0027, accessed February 1, 2006, at http://www.epa.gov/oppsrrd1/op/azm.htm. 

13. USEPA (U.S. Environmental Protection Agency), 2000, 

EFED risk assessment for reregistration eligibility science chapter for chlorpyrifos, fate and environmental 

risk assessment chapter: U.S. Environmental Protection 

Agency June 2000, accessed September 21, 2005, at http://

www.epa.gov/oppsrrd1/op/chlorpyrifos/efedrra1.pdf.

14. USEPA (U.S. Environmental Protection Agency), 2002, 

Interim reregistration eligibility decision for chlorpyrifos: 

U.S. Environmental Protection Agency, Prevention, Pesticides and Toxic Substances, EPA 738-R-01-007 February 

2002, accessed September 21, 2005, at http://www.epa.

gov/oppsrrd1/REDs/chlorpyrifos_ired.pdf.

15. USEPA (U.S. Environmental Protection Agency), 2000, 

Environmental risk assessment for diazinon: U.S Environmental Protection Agency, accessed September 28, 

2005, at http://www.epa.gov/pesticides/op/diazinon/risk_

oct2000.pdf.

16. USEPA (U.S. Environmental Protection Agency), 2004, 

Interim reregistration eligibility decision diazinon: U.S. 

Environmental Protection Agency, Prevention, Pesticides 

and Toxic Substances, EPA 738-R-04-006 May 2004, 

accessed September 21, 2005, at http://www.epa.gov/

oppsrrd1/REDs/diazinon_ired.pdf.

Appendix 3 165
17. USEPA (U.S. Environmental Protection Agency), 2000, 

Second revised environmental fate and effects assessment [disulfoton]: U.S. Environmental Protection Agency, 

Environmental Fate and Effects Division August 26, 2000, 

accessed October 7, 2005, at http://www.epa.gov/oppsrrd1/

op/disulfoton/EFED%20Chapter%208_26_00.pdf.

18. USEPA (U.S. Environmental Protection Agency), 2000, 

Malathion reregistration eligibility document, environmental fate and effects chapter, revised November 9, 

2000: U.S. Environmental Protection Agency, accessed 

September 22, 2005, at http://www.epa.gov/oppsrrd1/op/

malathion/efedrra.pdf.

19. USEPA (U.S. Environmental Protection Agency), 1999, 

EFED RED chapter for ethyl parathion (revised September 13, 1999): U.S. Environmental Protection Agency, 

accessed October 11, 2005, at http://www.epa.gov/

oppsrrd1/op/ethyl-parathion/efedrra.pdf.

20. USEPA (U.S. Environmental Protection Agency), 2003, 

Interim reregistration eligibility decision for methyl 

parathion: U.S. Environmental Protection Agency, Prevention, Pesticides and Toxic Substances, EPA 738-R-01-007 

February 2002, accessed October 12, 2005, at http://www.

epa.gov/oppsrrd1/REDs/methylparathion_ired.pdf.

21. USEPA (U.S. Environmental Protection Agency), 1998, 

Preliminary environmental fate and effects assessment, 

EFED science chapter for phorate RED August 7, 1998: 

U.S. Environmental Protection Agency, Environmental 

Fate and Effects Division, accessed October 20, 2005, at 

http://www.epa.gov/oppsrrd1/op/phorate/phorefed.pdf.

22. USEPA (U.S. Environmental Protection Agency), 1999, 

Revised environmental fate and effects assessment, 

outline of section C, Environmental assessment [phorate] September 14, 1999: U.S. Environmental Protection Agency, http://www.epa.gov/oppsrrd1/op/phorate/

phorateefedrevrisk2of6.pdf.

23. USEPA (U.S. Environmental Protection Agency), 1999, 

Revised environmental fate and effects assessment [terbufos], section 4, Ecological toxicity data, September 14, 

1999: U.S. Environmental Protection Agency, accessed 

October 4, 2005, at http://www.epa.gov/pesticides/op/

terbufos/efed_toxdata.pdf.

24. USEPA (U.S. Environmental Protection Agency), 2000, 

Environmental Fate and Effects Division science chapter for reregistration eligibility decision for propargite, 

Appendix 2, June 7, 2000: U.S. Environmental Protection 

Agency, Office of Pesticide Programs, Environmental Fate 

and Effects Division, accessed October 20, 2005, at http://

www.epa.gov/oppsrrd1/reregistration/propargite/efedrisk.

pdf.

25. USEPA (U.S. Environmental Protection Agency), 2003, 

Interim reregistration eligibility decision for atrazine, case 

no. 0062: U.S. Environmental Protection Agency, Office 

of Prevention, Pesticides and Toxic Substances, Office of 

Pesticide Programs, Special Review and Reregistration 

Division, accessed September 8, 2005, at http://www.epa.

gov/oppsrrd1/REDs/atrazine_ired.pdf.

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Atrazine MOA Ecological Subgroup—recommendations for aquatic community level of concern (LOC) and 

method to apply LOC(s) to monitoring data, final report, 

October 22, 2003: U.S. Environmental Protection Agency, 

Public Docket Number EPA-HQ-OPP-2003-0367-0007, 

accessed January 24, 2006, at http://www.regulations.gov.

27. USEPA (U.S. Environmental Protection Agency), 2003, 

Reregistration eligibility decision for diuron, list A, case 

0046: U.S. Environmental Protection Agency, Office of 

Prevention, Pesticides and Toxic Substances, Special 

Review and Reregistration Division, accessed September 

21, 2005, at http://www.epa.gov/oppsrrd1/REDs/diuron_

RED.pdf.

28. MacDonald, D.D., Ingersoll, C.G., and Berger, T.A., 2000, 

Development and evaluation of consensus-based sediment 

quality guidelines for freshwater ecosystems: Archives of 

Environmental Contamination and Toxicology, v. 39, p. 

20–31.

29. USEPA (U.S. Environmental Protection Agency), 2003, 

Procedures for the derivation of equilibrium partitioning 

sediment benchmarks (ESBs) for the protection of benthic 

organisms—dieldrin. U.S. Environmental Protection 

Agency, Office of Research and Development, EPA-600-

R-02-010, August 2003, accessed June 9, 2005, at http://

www.epa.gov/nheerl/publications/files/dieldrin.pdf.

30. USEPA (U.S. Environmental Protection Agency), 2003, 

Procedures for the derivation of equilibrium partitioning 

sediment benchmarks (ESBs) for the protection of benthic organisms—endrin: U.S. Environmental Protection 

Agency, Office of Research and Development, EPA-600-

R-02-009, August 2003, accessed June 9, 2005, at w.

31. USEPA (U.S. Environmental Protection Agency), 2003, 

Procedures for the derivation of equilibrium partitioning 

sediment benchmarks (ESBs) for the protection of benthic 

organisms—nonionics compendium: U.S. Environmental 

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Toxicological benchmarks for wildlife, 1996 revision: prepared by the Risk Assessment Program, Health Sciences 

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166 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
33. Newell, A.J., Johnson, D.W., and Allen, L.K., 1987, 

Niagara River Biota Contamination Project—fish flesh 

criteria for piscivorous wildlife: New York State Depart-

ment of Environmental Conservation, Division of Fish and 

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protection of wildlife that consume aquatic biota—DDT 

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1999: Winnipeg, Canadian Council of Ministers of the 

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35. CCME (Canadian Council of Ministers of the Environ-

ment), 1999, Canadian tissue residue guidelines for the 

protection of wildlife that consume aquatic biota—toxa-

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ards to fish, wildlife, and invertebrates—a synoptic review:

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and invertebrates—a synoptic review: U.S. Department of 

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85(1.21), Contaminant Hazard Reviews, Report 21, 49 p.

Appendix 3 167
Water: Human-Health Benchmarks

1. Pesticide compounds without human-health 

benchmarks:

The following pesticide compounds measured by NAWQA 

in stream water and ground water had no human-health 

benchmarks (names in parentheses are synonyms):

Acetochlor

Azinphos-methyl (Guthion)

Benfluralin

Bromoxynil

Chloramben methyl ester

Clopyralid

Dacthal mono-acid

2,4-DB

Deethylatrazine

Dichlobenil

Dichlorprop

2,6-Diethylaniline

DNOC

EPTC

Ethalfluralin

Ethoprop (Ethoprophos)

Fenuron

3-Hydroxycarbofuran

Linuron

MCPB

Methiocarb

Molinate

Napropamide

Neburon

Norflurazon

Oryzalin

Parathion (Ethyl parathion)

Pebulate 

Pendimethalin

cis-Permethrin

Phorate

Propanil

Propargite

Thiobencarb

Triallate

Triclopyr

2. Pesticide compounds with human-health 

benchmarks, but no exceedances:

The following pesticide compounds measured in water had 

human-health benchmarks available, but these benchmarks 

were never exceeded (names in parentheses are synonyms):

Acifluorfen

Alachlor

Aldicarb

Aldicarb sulfone

Aldicarb sulfoxide

Bentazon

Bromacil

Butylate

Carbaryl

Carbofuran

Chlorothalonil

Chlorpyrifos

2,4-D

Dacthal (DCPA)

p,p'-DDE

Dicamba

Disulfoton

Diuron

Fluometuron

Fonofos

alpha-HCH

Malathion

MCPA

Methomyl

Metolachlor

Metribuzin

Oxamyl

Parathion-methyl (Methyl parathion)

Picloram

Prometon

Pronamide (Propyzamide)

Propachlor

Propham

Propoxur (Baygon)

Simazine

2,4,5-T

Tebuthiuron

Terbacil

Terbufos

2,4,5-TP (Silvex)

Trifluralin

168 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Water: Aquatic-Life Benchmarks

1. Pesticide compounds without aquatic-life 

benchmarks:

The following pesticide compounds measured by NAWQA 

in stream water had no aquatic-life benchmarks (names in 

parentheses are synonyms):

Acetochlor

Chloramben methyl ester

Clopyralid

Cyanazine

Dacthal mono-acid

Deethylatrazine

Dicamba

Dichlorprop

2,6-Diethylaniline

Dinoseb

DNOC

Fenuron

Fonofos

alpha-HCH

3-Hydroxycarbofuran

MCPB

Neburon

Prometon

Propham

2,4,5-T

2,4,5-TP (Silvex)

2. Pesticide compounds with aquatic-life 

benchmarks, but no exceedances:

The following pesticide compounds measured in stream water 

had aquatic-life benchmarks available, but these benchmarks 

were never exceeded (names in parentheses are synonyms):

Acifluorfen

Aldicarb

Aldicarb sulfone

Aldicarb sulfoxide

Benfluralin

Bentazon

Bromacil

Bromoxynil

Butylate

Chlorothalonil

2,4-D

Dacthal (DCPA)

2,4-DB

Dichlobenil

EPTC

Ethalfluralin

Ethoprop (Ethoprophos)

Fluometuron

gamma-HCH (Lindane)

Linuron

MCPA

Methiocarb

Metolachlor

Metribuzin

Napropamide

Norflurazon

Oryzalin

Oxamyl

Pebulate

Pendimethalin

cis-Permethrin

Picloram

Pronamide (Propyzamide)

Propachlor

Propanil

Propoxur (Baygon)

Simazine

Tebuthiuron

Terbacil

Triallate

Trifluralin

Triclopyr

Appendix 3 169
Bed Sediment: Aquatic-Life Benchmarks

1. Pesticide compounds without aquatic-life 

benchmarks:

The following pesticide compounds measured by NAWQA in 

bed sediment had no aquatic-life benchmarks:

Aldrin

Chloroneb

Dacthal (DCPA)

alpha-HCH

beta-HCH

Heptachlor

Hexachlorobenzene

Isodrin

Mirex

Pentachloroanisole

cis-Permethrin

trans-Permethrin

Note: Several additional compounds measured, such as cischlordane and p,p'-DDE, did not have benchmarks themselves, 

but were part of pesticide groups (total chlordane and total 

DDT, respectively) that did have sediment benchmarks.

2. Pesticide compounds with aquatic-life 

benchmarks, but no exceedances:

All pesticide compounds or groups with benchmarks in 

sediment exceeded those benchmarks in one or more samples.

Whole Fish Tissue: Wildlife Benchmarks

1. Pesticide compounds without wildlife 

benchmarks:

The following pesticide compounds measured by NAWQA in 

whole fish had no benchmarks for fish-eating wildlife:

Dacthal (DCPA)

Pentachloroanisole

Note: Several additional compounds measured, such as cischlordane and p,p'-DDE, did not have benchmarks themselves, 

but were part of pesticide groups (total chlordane and total 

DDT, respectively) that did have wildlife benchmarks.

2. Pesticide compounds with wildlife 

benchmarks, but no exceedances:

The following pesticide compounds or groups measured 

in whole fish had wildlife benchmarks available, but these 

benchmarks were never exceeded (names in parentheses are 

synonyms):

beta-HCH

gamma-HCH (Lindane)

Total HCH

Heptachlor

Hexachlorobenzene

Total Methoxychlor

Mirex

170 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Appendix 4 171

Appendix 4— List of abbreviations, acronyms, and units of measurement

Note: Clarification or additional information is provided in parentheses.

ATSDR Agency for Toxic Substances and Disease Registry

AWQC-AL ambient water-quality criterion for the protection of aquatic life (USEPA)

CAS Chemical Abstract Service 

CCME Canadian Council of Ministers of the Environment

CERCLA Comprehensive Environmental Recovery, Compensation, and Liability Act

CWA Clean Water Act

DDD 1,1-dichloro-2,2-bis(chlorophenyl)ethane (DDT degradate)

DDE 1,1-dichloro-2,2-bis(chlorophenyl)ethylene (DDT degradate)

DDT 1,1,1-trichloro-2,2-bis(chlorophenyl)ethane

ESA ethanesulfonic acid

ESB equilibrium partitioning sediment benchmark (USEPA)

FFC fish flesh criterion for protection of fish-eating wildlife, noncancer values (NYSDEC)

FFDCA Federal Food, Drug, and Cosmetic Act

FIFRA Federal Insecticide, Fungicide, and Rodenticide Act

FQPA Food Quality Protection Act

ft foot (feet)

g gram

GCMS gas chromatography with mass spectrometric detection

HA-L lifetime health advisory (USEPA)

HPLC high performance liquid chromatography

IARC International Agency for Research on Cancer

kg kilogram

KH Henry’s law constant

Koc soil organic carbon-water partition coefficient

L liter

lb pound

LOAEC lowest-observed-adverse-effects concentration

LOC level of concern (USEPA)

MCL Maximum Contaminant Level (USEPA)

mg milligram

mi mile

mL milliliter

mm millimeter

NAWQA National Water-Quality Assessment Program (USGS)

NCBP National Contaminant Biomonitoring Program (U.S. Fish and Wildlife Service)

NLCD National Land Cover Data

NOAEC no-observed-adverse-effects concentration

NOAEL no-observed-adverse-effects level

NPDES National Pollutant Discharge Elimination System (USEPA)
Appendix 4— List of abbreviations, acronyms, and units of measurement

Note: Clarification or additional information is provided in parentheses.

NYSDEC New York State Department of Environmental Conservation

OPP Office of Pesticide Programs (USEPA)

OW Office of Water (USEPA)

ppb parts per billion

PTI pesticide toxicity index

SDWA Safe Drinking Water Act

TEC threshold effect concentration

TRG tissue residue guideline (Canadian)

TU tritium unit

USEPA U.S. Environmental Protection Agency

USGS U.S. Geological Survey

WET whole effluent toxicity (USEPA)

yr year

μg microgram

172 Pesticides in the Nation’s Streams and Ground Water, 1992–2001
Acknowledgments

Special thanks to the following individuals and organizations for their contributions:

NAWQA personnel in Study Units across the Nation for their contributions of data, research results, and 

reviews of findings

NAWQA’s many partners in the governmental and nongovernmental sectors who have helped guide 

scientific efforts and ensure that NAWQA information meets the needs of local, State, Tribal, regional, 

and national stakeholders

U.S. Environmental Protection Agency’s Office of Pesticide Programs and Office of Water for contributions 

of background information and data regarding current regulatory procedures and risk-assessment results

Reviewers:

Erica Michaels Brown, Association of Metropolitan Water Agencies

Judy Campbell-Bird, Environmental Policy and Communications Consulting

Claudia Copeland, Congressional Research Service

Rebeckah Freeman and Don Parrish, American Farm Bureau Federation

Jane Houlihan, Environmental Working Group

Debra Knopman, RAND Corporation

Timothy Kubiak and George Noguchi, U.S. Fish and Wildlife Service

Jane Nogaki and Paul Schwartz, Clean Water Action

Robin O’Malley, H. John Heinz III Center for Science, Economics and the Environment

Christine Reimer, National Ground Water Association

Marc Ribaudo, U.S. Department of Agriculture

Gary Rowe, U.S. Geological Survey

George Sabbagh, Bayer CropScience

Mark Sandstrom, U.S. Geological Survey

William Wilber, U.S. Geological Survey

Office of Pesticide Programs, U.S. Environmental Protection Agency

Office of Water, U.S. Environmental Protection Agency

Technical editors:

Sandra Cooper, U.S. Geological Survey

Chester Zenone, Editorial Consultant

Graphic design and layout: 

Yvonne Roque and Phillip Redman, U.S. Geological Survey

All photos by USGS personnel unless otherwise noted
Gilliom and others—The Quality of Our Nation’s Waters—Pesticides in the Nation’s Streams and Ground Water, 1992–2001—U.S. Geological Survey Circular 1291

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