United States

Department of

Agriculture Drinking Water from Forests

Forest Service and Grasslands

A Synthesis of the Scientific Literature

Southern

Research Station

General Technical

Report SRS-39

George E. Dissmeyer, Editor
The Editor:

George E. Dissmeyer, Water Quality Management—

Nonpoint Source Consultant, 145 Lake Forest Drive, NE,

Eatonton, GA 31024.

Cover photo: Hendersonville Reservoir Dam, Pisgah National Forest, North Carolina. Photo by Bill Lea.

_____________________________________________________________________________________________

The use of trade or firm names in this publication is for reader information and does not imply endorsement of any product or

service by the U.S. Department of Agriculture or other organizations represented here.

_____________________________________________________________________________________________

This publication reports research involving pesticides. It does not contain recommendations for their use, nor does it imply that

the uses discussed here have been registered. All uses of pesticides must be registered by appropriate State and/or Federal

agencies before they can be recommended.

CAUTION: Pesticides can be injurious to humans, domestic animals, desirable plants, and fish or other wildlife—if they are

not handled or applied properly. Use all herbicides selectively and carefully. Follow recommended practices for the disposal of

surplus pesticides and their containers.

September 2000

USDA Forest Service

Southern Research Station

Asheville, North Carolina
i

Drinking Water from Forests

and Grasslands

A Synthesis of the Scientific Literature

George E. Dissmeyer, Editor

“after refreshing ourselves we proceeded on to the top of the dividing

ridge from which I discovered immence ranges of high mountains still to

the West of us with their tops partially covered with snow. I now

decended the mountain about 3/4 of a mile which I found much steeper

than on the opposite side, to a handsome bold runing Creek of cold Clear

water. here I first tasted the water of the great Columbia river.”

—from Meriwether Lewis’ journal, August 12, 1805
iii

Table of Contents

Page

Executive Summary ......................................................................................................................................... ix

Part I: Introduction

Chapter 1: Goals of this Report ...................................................................................................................... 3

Douglas F. Ryan and Stephen Glasser

The Importance of Safe Public Drinking Water ................................................................................................. 3

Drinking Water from Forests and Grasslands ..................................................................................................... 3

The Purpose and Scope of this Document .......................................................................................................... 4

How to Use this Document ................................................................................................................................ 4

Acknowledgments .............................................................................................................................................. 6

Literature Cited ................................................................................................................................................... 6

Chapter 2: Drinking Water Quality ................................................................................................................ 7

F.N. Scatena

Introduction ........................................................................................................................................................ 7

Chemical Properties ........................................................................................................................................... 7

Physical Properties ............................................................................................................................................. 20

Biological Properties .......................................................................................................................................... 22

Literature Cited ................................................................................................................................................... 25

Chapter 3: Watershed Processes—Fluxes of Water, Dissolved Constituents, and Sediment .................... 26

F.J. Swanson, F.N. Scatena, G.E. Dissmeyer, M.E. Fenn, E.S. Verry, and J.A. Lynch

Introduction ........................................................................................................................................................ 26

The Integrated Hydrologic System ..................................................................................................................... 26

Effects of Nitrogen Deposition on Stream and Ground Water Quality .............................................................. 32

Sediment Production and Transport ................................................................................................................... 34

North Santiam River Case Study ........................................................................................................................ 35

Natural Disturbance Processes ........................................................................................................................... 36

Cumulative Watershed Effects ........................................................................................................................... 37

Management and Policy Considerations ............................................................................................................ 37

Research Needs .................................................................................................................................................. 38

Key Points .......................................................................................................................................................... 39

Literature Cited ................................................................................................................................................... 39

Chapter 4: Economic Issues for Watersheds Supplying Drinking Water ................................................... 42

Thomas C. Brown

Introduction ........................................................................................................................................................ 42

Cost Minimization .............................................................................................................................................. 43

Opportunities for Cost Savings .......................................................................................................................... 45

Complexity ......................................................................................................................................................... 47

Cost Savings from Targeting Upstream Control Efforts .................................................................................... 48

Bringing About an Efficient Cost Allocation ..................................................................................................... 48

Conclusion .......................................................................................................................................................... 50

Literature Cited ................................................................................................................................................... 50
iv

Part II: Effects of Recreation and the Built Environment on Water Quality

Page

Chapter 5: Hydromodifications—Dams, Diversions, Return Flows, and Other Alterations

of Natural Water Flows .................................................................................................................................... 55

Stephen P. Glasser

Introduction ........................................................................................................................................................ 55

Effects of Dams and Impoundments on Water Quality ...................................................................................... 56

Water Diversion Structures and Water Import/Export Between Watersheds ..................................................... 57

Water Well Effects on Drinking Water Quality .................................................................................................. 58

Sewage Effluent and Sludge/Biosolids Applications to Forest and Rangeland ................................................. 58

Wetland Drainage ............................................................................................................................................... 59

Reclaimed Water and Return Flows ................................................................................................................... 59

Reliability and Limitations of Findings .............................................................................................................. 59

Research Needs .................................................................................................................................................. 60

Key Points .......................................................................................................................................................... 60

Literature Cited ................................................................................................................................................... 60

Chapter 6: Urbanization .................................................................................................................................. 62

Wayne C. Zipperer, Karen Solari, and Beverly A. Young

Introduction ........................................................................................................................................................ 62

Issues and Risks .................................................................................................................................................. 62

Wastewater Treatment ........................................................................................................................................ 64

Urban Runoff ...................................................................................................................................................... 65

Underground Storage Tanks ............................................................................................................................... 67

Abandoned Wells ................................................................................................................................................ 68

Solid Waste Landfills and Other Past Land Uses ............................................................................................... 70

Literature Cited ................................................................................................................................................... 72

Chapter 7: Concentrated Recreation .............................................................................................................. 74

Myriam Ibarra and Wayne C. Zipperer

Introduction ........................................................................................................................................................ 74

Campgrounds ...................................................................................................................................................... 74

Water Recreation ................................................................................................................................................ 75

Winter Recreation ............................................................................................................................................... 77

Increased Traffic ................................................................................................................................................. 78

Literature Cited ................................................................................................................................................... 79

Chapter 8: Dispersed Recreation .................................................................................................................... 81

David Cole

Introduction ........................................................................................................................................................ 81

Issues and Risks .................................................................................................................................................. 81

Findings from Studies ........................................................................................................................................ 81

Reliability and Limitation of Findings ............................................................................................................... 83

Research Needs .................................................................................................................................................. 83

Key Points .......................................................................................................................................................... 83

Literature Cited ................................................................................................................................................... 84

Chapter 9: Roads and Other Corridors ......................................................................................................... 85

W.J. Elliot

Introduction ........................................................................................................................................................ 85

Altered Hydrology .............................................................................................................................................. 86

Sedimentation ..................................................................................................................................................... 86

Hydrocarbons, Cations, and Related Pollutants ................................................................................................. 92

Fuels and Other Contaminants from Accidental Spills ...................................................................................... 95

Pipeline Failures ................................................................................................................................................. 95

Literature Cited ................................................................................................................................................... 97
v

Part III: Effects of Vegetation Management on Water Quality

 Page

Chapter 10: Timber Management .................................................................................................................. 103

John D. Stednick

Introduction ........................................................................................................................................................ 103

Erosion/Sedimentation ....................................................................................................................................... 103

Stream Temperature ............................................................................................................................................ 108

Nutrients ............................................................................................................................................................. 110

Fertilizer ............................................................................................................................................................. 113

Literature Cited ................................................................................................................................................... 115

........................................................................................................................................................................

Chapter 11: Forest Succession ......................................................................................................................... 120

Wayne Swank

Introduction ........................................................................................................................................................ 120

Nutrients ............................................................................................................................................................. 120

Sediment ............................................................................................................................................................. 122

Literature Cited ................................................................................................................................................... 123

........................................................................................................................................................................

Chapter 12: Fire Management ........................................................................................................................ 124

Johanna D. Landsberg and Arthur R. Tiedemann

Introduction ........................................................................................................................................................ 124

Sediment and Turbidity ...................................................................................................................................... 124

Temperature ........................................................................................................................................................ 128

Chemical Water Quality ..................................................................................................................................... 128

Nitrogen .............................................................................................................................................................. 129

Phosphorus ......................................................................................................................................................... 132

Sulfur .................................................................................................................................................................. 132

Chloride .............................................................................................................................................................. 132

Total Dissolved Solids ........................................................................................................................................ 132

Trace Elements ................................................................................................................................................... 132

Effects on Ground Water .................................................................................................................................... 132

Literature Cited ................................................................................................................................................... 135

........................................................................................................................................................................

Chapter 13: Pesticides ..................................................................................................................................... 139

J.L. Michael

Introduction ........................................................................................................................................................ 139

Issues and Risks, Pesticide Application ............................................................................................................. 139

Issues and Risks, Toxicity .................................................................................................................................. 140

Findings from Studies ........................................................................................................................................ 143

Reliability and Limitation of Findings ............................................................................................................... 147

Research Needs .................................................................................................................................................. 148

Key Points .......................................................................................................................................................... 149

Literature Cited ................................................................................................................................................... 149
vi

Part IV: Effects of Grazing Animals, Birds, and Fish on Water Quality

Page

Chapter 14: Domestic Grazing ........................................................................................................................ 153

John C. Buckhouse

Introduction ........................................................................................................................................................ 153

Erosion and Sedimentation ................................................................................................................................. 153

Bacteria and Protozoa ......................................................................................................................................... 154

Chemical and Nutrient Impacts .......................................................................................................................... 155

Literature Cited ................................................................................................................................................... 156

Chapter 15: Wildlife ......................................................................................................................................... 158

Arthur R. Tiedemann

Introduction ........................................................................................................................................................ 158

Issues and Risks .................................................................................................................................................. 158

Findings from Studies ........................................................................................................................................ 160

Reliability and Limitation of Findings ............................................................................................................... 160

Research Needs .................................................................................................................................................. 161

Key Points .......................................................................................................................................................... 162

Literature Cited .................................................................................................................................... 162

Chapter 16: Water Birds.................................................................................................................................. 164

Christopher A. Nadareski

Introduction ........................................................................................................................................................ 164

Contamination .................................................................................................................................................... 164

Water Birds as Vectors of Contamination ........................................................................................................... 166

Seasonality of Impacts ........................................................................................................................................ 166

Reliability and Limitations of Findings .............................................................................................................. 166

Key Points .......................................................................................................................................................... 167

Case Study: New York City Waterfowl Management Program ......................................................................... 167

Literature Cited ................................................................................................................................................... 167

Chapter 17: Fish and Aquatic Organisms ..................................................................................................... 169

C. Andrew Dolloff

Introduction ........................................................................................................................................................ 169

Fish Hatcheries and Aquaculture Facilities ........................................................................................................ 169

Chemical Reclamation........................................................................................................................................ 171

Restoration and Reintroduction of Populations and Communities .................................................................... 172

Physical Habitat .................................................................................................................................................. 173

Liming of Acidified Waters ................................................................................................................................ 174

Literature Cited ................................................................................................................................................... 174
vii

Part V: Effects of Mining and Oil and Gas Development on Water Quality

Page

Chapter 18: Hardrock Mining ........................................................................................................................ 179

Mike Wireman

Introduction ........................................................................................................................................................ 179

Mining Methods ................................................................................................................................................. 179

Ore Processing .................................................................................................................................................... 180

Water Management ............................................................................................................................................. 181

Waste Management ............................................................................................................................................ 181

Mine Closure ...................................................................................................................................................... 181

Issues and Risks.................................................................................................................................................. 182

Erosion and Sedimentation ................................................................................................................................. 182

Acid Rock Drainage ........................................................................................................................................... 183

Cyanide Leaching ............................................................................................................................................... 183

Transport of Dissolved Contaminants ................................................................................................................ 184

Findings from Studies ........................................................................................................................................ 184

Reliability and Limitations of Findings.............................................................................................................. 185

Research Needs .................................................................................................................................................. 185

Key Points .......................................................................................................................................................... 186

Literature Cited ................................................................................................................................................... 186

Chapter 19: Coal Mining ................................................................................................................................. 187

Mike Wireman

Introduction ........................................................................................................................................................ 187

Mining Methods ................................................................................................................................................. 187

Coal Preparation ................................................................................................................................................. 187

Waste Management ............................................................................................................................................ 187

Environmental Regulation .................................................................................................................................. 188

Issues and Risks.................................................................................................................................................. 188

Findings from Studies ........................................................................................................................................ 188

Reliability and Limitations of Findings.............................................................................................................. 188

Research Needs .................................................................................................................................................. 188

Key Points .......................................................................................................................................................... 189

Literature Cited ................................................................................................................................................... 189

Chapter 20: Oil and Gas Development .......................................................................................................... 190

R.J. Gauthier-Warinner

Introduction ........................................................................................................................................................ 190

Issues and Risks.................................................................................................................................................. 191

Findings from Studies ........................................................................................................................................ 193

Reliability and Limitation of Findings ............................................................................................................... 193

Research Need .................................................................................................................................................... 193

Key Points .......................................................................................................................................................... 193

Literature Cited ................................................................................................................................................... 194
viii

Part VI: Implications for Source Water Assessments and for Land

Management and Policy

Page

Chapter 21: Future Trends and Research Needs in Managing Forests and Grasslands

as Drinking Water Sources ............................................................................................................................... 197

F.N. Scatena

Introduction ......................................................................................................................................................... 197

Environmental Change ........................................................................................................................................ 197

Technological Change ......................................................................................................................................... 198

Administrative Change ........................................................................................................................................ 198

Site-Specific Considerations ............................................................................................................................... 199

Conclusion ........................................................................................................................................................... 200

Literature Cited .................................................................................................................................................... 200

Chapter 22: Synthesis........................................................................................................................................ 202

Douglas F. Ryan

Introduction ......................................................................................................................................................... 202

Drinking Water Contaminants and Treatments ................................................................................................... 202

Cumulative Effects .............................................................................................................................................. 202

Effects of Natural Processes and Human Activities ............................................................................................ 203

Implications of Scientific Uncertainty................................................................................................................. 204

Implications for Source Water Assessments ........................................................................................................ 205

Source Water Protection as a Priority for Land Management ............................................................................. 205

Literature Cited .................................................................................................................................................... 206

Part VII: Appendices

Appendix A: City of Baltimore Municipal Reservoirs, Incorporating Forest

Management Principles and Practices............................................................................................................. 209

Robert J. Northrop

Appendix B: Managing the Shift from Water Yield to Water Quality on Boston’s

Water Supply Watersheds................................................................................................................................. 212

Thom Kyker-Snowman

Appendix C: Cumulative Impacts of Land Use on Water Quality in a Southern

Appalachian Watershed .................................................................................................................................... 215

Wayne T. Swank and Paul V. Bolstad

Appendix D: Protozoan Pathogens Giardia and Cryptosporidium ................................................................ 218

David Stern

Appendix E: Water Treatment Technologies Tables ...................................................................................... 225

Gary Logsdon

List of Figures .................................................................................................................................................... 230

List of Tables ...................................................................................................................................................... 232

Glossary of Abbreviations and Acronyms ....................................................................................................... 234

Glossary of Terms .............................................................................................................................................. 237

Subject Index ..................................................................................................................................................... 242
Part I:

Introduction
ix

Executive Summary

The Safe Drinking Water Act Amendments of 1996 require every State to perform source water assessments of all public

drinking water sources and make the results public by 2003. Forests and grasslands serve as sources of many public drinking

water supplies, and managers of these lands are expected to participate in preparing assessments and to work with the public to

assure safe drinking water. To help managers of forests and grasslands meet this requirement, this report reviews the current

scientific literature about the potential of common land-use practices to introduce contaminants that pose risks to human health

into public drinking water sources. Potential audiences for this report include managers of national forests and grasslands and

managers of other public and private lands with similar uses. Operators of public drinking water utilities and citizens’ groups

concerned with drinking water may also find this report useful.

Safe drinking water is essential to protect public health. Modern drinking water treatment can reduce most contaminants in

source water to acceptable levels before it is delivered to consumers, but costs increase significantly when more rigorous

treatment is needed to cleanse contaminated source water. Managing land to prevent source water contamination may be more

cost-effective and may better protect human health than treating water after it has been contaminated.

Water from forests and grasslands is usually cleaner than water from urban and agricultural areas. Nevertheless, many common

practices on forests and grasslands can contaminate drinking water sources. Soil disturbing activities such as road construction

and maintenance, forest harvesting, and intermixed urban and wildland uses can introduce sediment into drinking water sources.

Disease organisms may enter source waters from: (1) recreation and other human activities that lack developed sanitary facilities, (2) malfunctioning sewage disposal facilities, and (3) wild and domestic animals concentrated near source waters. Nutrients

may enter source water from fertilizer and from atmospheric deposition of nitrogen compounds. Toxic chemicals may reach

source water from pest control; from extraction of minerals, oil, and gas; from accidental chemical spills along highways and

utility corridors; and from leaking underground storage tanks.

Gaps exist in the scientific understanding of the effects of many land-use practices on drinking water sources. For example,

pathogens in wild animal populations and their transmission to source water are poorly known. Risk of contamination from

recreation that occurs in areas without developed sanitary facilities is largely unstudied. Effects of multiple land uses that

overlap in time and space across large watersheds are difficult to predict with current knowledge. Managers should consider

uncertainties due to these unknowns in land-use decisions until research fills these knowledge gaps.

Source water assessments for forest and grassland watersheds are not likely to be fundamentally different from those in areas

with other land uses. Scientific information will need to be applied locally on a case-by-case basis to consider what natural and

human activities have a reasonable potential to introduce contaminants that are likely to reach a drinking water intake. Assessments will need to integrate across conventional disciplinary boundaries to assess the overall degree of risk to drinking water

sources. Scientists, land managers, and the public will need to cooperate to translate the basic information in this report into

meaningful source water assessments.

Keywords: Economics, land use, nutrients, pathogens, sediments, source water assessments, toxics.
3

Chapter 1

Goals of this Report

Douglas F. Ryan and Stephen Glasser1

U.S. Congress chose source water protection as a strategy

for ensuring safe drinking water because of its high potential

to be cost-effective. A poor source of water can substantially

increase the cost of treatment to make the water drinkable.

When source water is so contaminated that treatment is not

feasible, developing alternative water supplies can be

expensive and cause delays in providing safe, affordable

water. Delineating areas that supply water and inventorying

potential sources of contamination will help communities

know the threats to their drinking water. Communities can

then more effectively and efficiently address these threats.

Drinking Water from Forests and Grasslands

Forests and grasslands have long been relied upon as

sources of clean drinking water for two reasons: (1) forests

mainly grow under conditions that produce relatively

reliable water runoff, and (2) properly managed forests and

grasslands can yield water relatively low in contaminants

when compared with many urban and agricultural land uses.

We estimate that at least 3,400 towns and cities currently

depend on National Forest System watersheds for their

public water supplies. In addition, the national forests and

grasslands have over 3,000 public water supplies for

campgrounds, administrative centers, and similar facilities.

Communities that draw source water from national forests

and grasslands provide a public water supply to 60 million

people, or one-fourth of the people served by public water

supplies nationwide. Since 70 percent of the forest area in

the United States is outside of the National Forest System,

the number of people served by all forests and grasslands is

far greater.

With the large number of public water supplies on forests

and grasslands, there is a high likelihood that many forest

and grassland managers will be involved in the process of

planning, implementing, or reacting to public concerns

related to SWA’s. The level of involvement in this process

will probably vary from place to place depending on the

requirements of each State, the degree of public attention

that particular management activities receives, and the

potential of specific land uses to affect source waters. At the

time of writing this document, it is difficult to predict to

The Importance of Safe Public Drinking Water

The U.S. Congress justified passing the Safe Drinking Water

Act Amendments of 1996 (SDWA) (Public Law 104–182)

codified at 42 U.S.C. sec. 300j–14, by stating “safe drinking

water is essential to the protection of public health.” For

over 50 years, a basic axiom of public health protection has

been that safe drinking water reduces infectious disease and

extends life expectancy (American Water Works Association

1953). Although most U.S. residents take safe public

drinking water for granted, assuring its safety remains a high

national priority. Large investments are made by all levels of

government to maintain and upgrade our public water

systems.

To strengthen that process, the SDWA mandates that greater

protection and information be provided for the 240 million

Americans who are served by public water supplies. Section

1453 of the SDWA requires all States to complete source

water assessments (SWA’s) of their public drinking water

supplies by 2003. To meet this requirement, each State and

participating tribe will delineate the boundaries of areas that

serve as sources for individual public drinking water

systems, identify significant potential sources of contamination, and determine how susceptible each system is to

contamination. Source water assessments are required for all

public drinking water supplies regardless of the ownership

of the drinking water system or the land that comprises its

source area. Results of SWA’s will be made public and will

assist local planners, tribes, and Federal and State Governments to make more informed decisions to protect drinking

water sources.

To get information about a source water assessment program

(SWAP) from a particular State, go to the U.S. Environmental Protection Agency (EPA) homepage to view the SWAP

contact list. This site includes names and telephone numbers

of State source water contacts and hotlinks to existing State

homepages for more information. The EPA homepage can

be found at http://epa.gov/OGWDW/protect.html.

1 Staff Watershed Specialist, Wildlife, Fish, Water, and Air Research Staff;

and Water Rights and Uses Program Manager, Watershed and Air

Management, USDA Forest Service, Washington, DC, respectively.
4

what degree particular managers may become involved with

this process. We have assembled current scientific knowledge in a useful form that will help managers protect the

safety of drinking water sources and be better-informed

participants in SWA’s.

The Purpose and Scope of this Document

This document was written to assist forest and grassland

managers in their efforts to comply with the SDWA by

providing them with a review and synthesis of the current

scientific literature about the effects of managing these lands

on public drinking water sources. This is not a decision

document. Its audience includes managers of national

forests and grasslands as well as managers of public and

private forests and grasslands. Managers of public water

supplies and community groups concerned with drinking

water may also find this document useful.

This report’s focus is restricted to potential contamination of

source water associated with ordinary land uses in national

forests and grasslands. It does not treat the delineation of

source areas because the EPA and the States will decide

those criteria. We chose conventional land uses on national

forests and grasslands because they clearly come under the

mandate of the U.S. Department of Agriculture, Forest

Service (Forest Service), the principal sponsor of this

document, and because a significant portion of the public

depends on national forests and grasslands for water. We did

include grazing and land uses that occur where urban areas

border on or intermix with forests and grasslands. The report

does not address large urban developments, large industrial

complexes, row crop agriculture, or concentrated animal

feeding operations because they come more appropriately

under the oversight of other agencies. We focus on issues for

public water supplies, rather than those of small, private

water sources for individual families, because only public

supplies are examined in SWA’s.

The processes reviewed in this report occur at spatial scales

ranging from a few square yards (meters) to many millions

of acres (hectares). Most scientific studies, however, have

been done at relatively small scales. Inferences about larger

areas are drawn mostly from models or extrapolations based

on those small-scale studies. Where regional differences in

effects of land management were reported in the literature,

the authors indicated them in this document. If not, we did

not make regional distinctions. Several conventions are used

by the scientific and land management communities for

classifying geographic, climatic, and ecological zones with

similar characteristics into ecoregions, but no standard

system of classification has been endorsed across relevant

scientific disciplines or Federal Agencies. For this reason,

we cited whatever ecoregions were used in the literature.

How to Use this Document

This document is intended to be used by managers as a

reference for assessing watersheds and planning programs to

minimize the effects of land management practices on the

quality of drinking water sources. When managers are

concerned with the potential of a particular land management practice, they can consult the chapter summarizing

what is known about the effects of that practice. Managers

should note both what is known and what is not known from

scientific studies. Known information may provide a means

to estimate the effects of a particular practice. What is

unknown is equally important because it may indicate which

management actions entail risk because their effects are not

well understood.

We wish to emphasize the importance of using scientific

information as a basis for management. Managers often are

forced by circumstances to make decisions based on

incomplete knowledge. They compensate by filling information gaps with reasonable assumptions. Each such assumption carries the risk of unintended consequences. Use of

scientific data in decision-making has the advantage that

many of the important conditions that affect outcomes have

been controlled or measured, and critical assumptions are

often carefully spelled out. When decisions are based on

anecdotal experience, less may be known about conditions

that affect outcomes, and key assumptions about these

conditions may not be explicit. Decisions that draw on

scientific information, therefore, reduce the risk of unexpected outcomes.

The subjects covered are broadly and briefly summarized.

When managers need to go more deeply into a topic, they

should use the scientific literature that is cited in each

chapter as an entry point into the larger body of knowledge

that underlies each of the chapters. Wherever possible, the

scientific information that is cited has been peer reviewed

and published. Case studies presented are meant to illustrate

the complexity of actual management situations and are not

necessarily based on peer-reviewed literature.

To synthesize the scientific information into a form that

answers questions relevant to managers required that the

authors use their best professional judgement both to draw

together diverse sources and to evaluate their validity.

Exercising this judgement is necessary to make this document more useful than a mere compilation of data or

annotated bibliography. We have made every effort to make

Goals of this Report
5

apparent the distinction between published scientific

observations and logical synthesis on the part of the authors.

This document has undergone a rigorous peer review by

professional scientists and managers from inside and outside

government to critique the validity and currency of its

sources, syntheses, and conclusions. The finished document

has been revised to consider and respond to the comments of

these reviewers.

Although this document is separated into chapters by types

of land use, we recognize that in most practical situations

effects on source waters result from the cumulative effects of

multiple land uses that often overlap in space and change

over time. To address this issue we direct readers to chapter

2, which covers the natural processes of watersheds that

overlay all land uses, and to chapter 3, which summarizes

the cumulative effects of multiple land uses distributed over

space and time.

In this document we concentrate on issues that arise from the

need of managers to comply with the SDWA. This is only

one of the many policies and laws that currently govern the

actions of national forest and grassland managers. A provision of the Organic Act of 1897 (30 Stat. 11), codified at 16

U.S.C. Subsec. 473–475, 477–482, 551, that established the

national forests “for the purpose of securing favorable

conditions of water flows,” has been interpreted to authorize

managing this land for water resources. Administration of

national forests is currently guided primarily by four laws:

(1) the Multiple Use-Sustained Yield Act (Public Law 86–

517), codified at 16 U.S.C. sec.525 et seq.; (2) the National

Environmental Policy Act (Public Law 91–190), codified at

16 U.S.C. sec.4321 et seq.; (3) the Forest and Rangeland

Renewable Resources Planning Act (Public Law 93–378),

codified at 16 U.S.C. sec.1600 et seq.; and (4) the National

Forest Management Act (Public Law 94–588). Forest and

grassland managers also must comply with many environmental statutes including the Endangered Species Act

(Public Law 93–205), codified at 16 U.S.C. sec.1531 et seq.;

the Clean Water Act (Public Law 80–845), codified at 33

U.S.C. Sec.1251; and the Clean Air Act (Public Law 84–

159), codified at 42 U.S.C. sec.7401 et seq. Activities of the

Forest Service with State and private landowners were

authorized by the Cooperative Forestry Assistance Act

(Public Law 95–313) and amended in the 1990 Farm Bill

(Public Law 101–624), codified at 16 U.S.C. Subsec. 582a,

582a–8, 1648, 1642 (note), 1647a, 2101 (note), 2106a, 2112

(note), 6601 (note). The Forest and Rangeland Renewable

Resources Act (Public Law 93–378), with amendments in

the 1990 Farm Bill (Public Law 101–624), provided authority for research by the Forest Service. For a more complete

listing of relevant laws and the text of these laws, see U.S.

Department of Agriculture, Forest Service (1993). Over

time, the laws and policies that guide public land use have

evolved in response to changes in perceived public needs

and will probably continue to change in the future.

A number of laws that affect forest and grassland management require the use of best management practices (BMP’s).

These practices vary widely in their application and effectiveness from State to State and continually evolve in

response to new environmental concerns, technology, and

scientific evidence (Dissmeyer 1994). This document does

not cite or endorse specific BMP’s but rather presents

scientific evidence that has the potential to serve as a basis

for developing practices that more effectively protect source

water.

Some laws and prudent practice require that environmental

monitoring be used to assess the outcomes of land management. We considered the broad topic of monitoring to be

beyond the scope of our effort, but implicit throughout this

document is the assumption that monitoring of outcomes

should be an integral part of land management. Scientific

evidence does not eliminate all risks of unforeseen outcomes, and where scientific studies are lacking, risks are

likely to be higher. Monitoring land-use practices will help

to protect public health and other important values.

This document focuses narrowly on protecting human health

by protecting drinking water. We acknowledge that managers must consider a much wider range of values in most

land-use decisions. It is not our intent to tell managers how

to weigh a spectrum of values or how to decide among

them. Rather we wish to inform managers about specific

effects on drinking water so that they can better take these

effects into consideration when they make land-use

decisions.

Chapter 1
6

Acknowledgments

We thank a number of organizations and individuals for

their support in developing this document. They include the

Research and Development, National Forest System, and

State and Private Forestry Deputy Areas of the U.S. Department of Agriculture, Forest Service; the U.S. Environmental

Protection Agency, Office of Ground Water and Drinking

Water; and the National Council for Air and Stream Improvement; each of which contributed funds and expertise

toward this effort. In addition, we acknowledge the contributions of the American Water Works Association, the

Centers for Disease Control and Prevention, the U.S.

Geological Survey, the State Forester of Massachusetts, and

numerous peer reviewers.

Goals of this Report

Literature Cited

American Water Works Association. 1953. Water quality and treatment.

New York: American Water Works Association, Inc. 451 p.

Dissmeyer, George E. 1994. Evaluating the effectiveness of forestry best

management practices in meeting water quality goals and standards.

Misc. Publ. 1520. Washington, DC: U.S. Department of Agriculture,

Forest Service. 166 p.

U.S. Department of Agriculture, Forest Service. 1993. The principal

laws relating to Forest Service activities. Washington, DC: U.S.

Government Printing Office. 1,163 p.
7

Chapter 2

Drinking Water Quality

F.N. Scatena1

treatment. Considerable treatment may be required to purify

water meeting the ambient standard to comply with the

drinking water standard. As effects on human health from

exposure to contaminants in drinking water become better

understood and as new substances are released to the

environment, changes in drinking water standards can be

expected in the future.

Chemical Properties

Water is formed by the covalent union of two hydrogen (H)

atoms and one oxygen (O) atom. These atoms are joined in

an unsymmetrical arrangement where the hydrogen end of

the molecule has a slight positive charge and the oxygen end

a slight negative charge. This arrangement of unbalanced

electrical charges creates the dipolar characteristic that gives

the molecule the remarkable ability to act as both an acid

and a base and be a solvent for cations, anions, and some

types of organic matter. This arrangement also allows water

molecules to form hydrogen bonds with adjacent water

molecules. These bonds are responsible for water’s high

viscosity, high cohesion and adhesion, high surface tension,

high melting and boiling points, and the large temperature

range through which it is a liquid.

As water travels across the landscape, it interacts with its

environment through a variety of chemical processes (table

2.5). In the process, it picks up and transports dissolved

gases, cations and anions, amorphous organics, trace metals,

and particulates. The most common positively charged ions,

or cations, include calcium (Ca+2), magnesium (Mg+2),

sodium (Na+1), potassium (K+1), and ammonium (NH4

+1).

The most common anions, or negatively charged ions,

include nitrate (NO3

-1), sulfate (SO4-2), chloride (Cl-1), and

several different forms of phosphorus (P). Most amorphous

substances are organic carbon-based compounds that readily

adsorb and exchange cations. Common particulates include

mineral particles, i.e., inorganic sediment, organic debris,

and microscopic organisms (plankton, diatoms, etc.). Both

the chemical behavior (table 2.6) and the origin of contamination (table 2.1) vary with the type of chemical

contaminants.

Introduction

Watersheds are topographically defined areas drained by

connecting stream channels that discharge water, sediment,

and dissolved materials through a common outlet. The term

is synonymous with drainage basin and catchment and can

refer to a large river basin or the area drained by a single

ephemeral stream. Watersheds are commonly classified by

physiography (headwater, steeplands, lowland, etc.),

environmental condition (pristine, degraded, etc.), or their

principal use or land cover (forest, urban, agricultural,

municipal water supply, etc.).

Municipal watersheds are managed to provide a sustainable

supply of high-quality, safe drinking water at minimum

environmental and economic costs. Many activities within a

watershed can contaminate water (table 2.1), and most

supplies are not suitable for human consumption without

some form of treatment. This chapter provides an overview

of the chemical and physical processes that affect the

chemistry and quality of water as it travels across the

landscape. The appendix presents information on treatment

techniques (appendix tables E.1–E.4) that are used for

controlling common contaminants (National Research

Council 1997).

Water quality is a relative concept that reflects measurable

physical, chemical, and biological characteristics in relation

to a specific use. The suitability of water for domestic use is

typically defined by taste, odor, color, and the abundance of

organic and inorganic substances that pose risks to human

health (table 2.2). In the United States, suitability is formally

defined in legally enforceable primary standards (table 2.3)

and in recommended or secondary guidelines (table 2.4).

The States will focus on the contaminants listed in tables 2.3

and 2.4 in their source water assessments.

Standards for drinking water apply to water that is delivered

to consumers after it has been treated to remove contaminants, but not to source water as it is withdrawn from

surface or ground water. Ambient standards set under the

Clean Water Act (Public Law 80–845) for streams or lakes

are not intended to ensure that water is drinkable without

1 Ecosystem Team Leader, USDA Forest Service, International Institute of

Tropical Forestry, Río Piedras, PR.
8

Drinking Water Quality

Table 2.1—Summary of common water pollutants by land-use activities

Land use and Spatial Major types Pollution

type of activity distribution of pollution indicators

Forests

Harvesting Diffuse N, O Sediment

Camping, hunting Diffuse FC, O, S FC, garbage

Skiing Diffuse, line N, I, S Salts, sediment

Rangeland

Grazing Diffuse FC, N, O NO3

-1, sediment

Urbanization

Unsewered sanitation Point, diffuse N, FC, O, S NO3

-1, NH4+1,

FC, DOC, ClLeaking sewers Point, line N, FC, O, S FC, NH4

+1, NO3-1

Leaking fuel tanks Point O HC, DOC

Storm drainage Line, diffuse I, H, O, S Cl-, sediment

Industrial

Leaking tanks Point O, S, H Variable, HC

Spills Point, diffuse O, S, H Variable

Aerial fallout Diffuse S, I, N, O SO4

-2, NO3-1, HC

Agriculture

Cropland Diffuse N, O, S, P NO3

-1, sediment

Livestock Point, diffuse FC, N, O NO3

-1, sediment

Mineral extraction Point, diffuse H, I Variable, sediment

DOC = dissolved organic carbon; FC = fecal coliform; H = heavy metals; HC = hydrocarbons; I = inorganic salts; N = nutrient;

NH4

+1 = ammonium; NO3-1 = nitrate; O = organic load; P = phosphorous; S = synthetic organic compounds; SO4-2 = sulfate.

Source: Updated from Foster and Gomes 1989.

Table 2.2—Common types of water contaminant guidelines for different water usesa

Human

Contaminant consumption Irrigation Livestock Fisheries Recreation

Coliform bacteria * *

Nematode eggs *

Particulate matter * *

Dissolved oxygen (BOD, COD) * *

Nitrates * * *

Nitrites * * *

Salinity * * * *

Inorganic pollutants (trace metals) * * * * *

Organic pollutants * * *

Pesticides * *

BOD = biological oxygen demand; COD = chemical oxygen demand.

a

An * indicates that guidelines typically exist for a particular use. The absence of an * indicates that no guidelines exist for a

particular use.

Source: Adapted from GEMS 1991.
9

Chapter 2

Table 2.3—National primary drinking water regulationsa (States are expected to focus attention on risks

related to the contaminants listed in their source water assessments.)

MCL Potential health effects Sources of contaminant

Contaminants MCLG or TT from ingestion of water in drinking water

- - - Milligrams per liter - - -

Inorganic chemicals

Antimony 0.006 0.006 Increase in blood cholesterol, Discharge from petroleum refineries,

decrease in blood glucose fire retardants, ceramics, electronics,

solder

Arsenic Noneb .05 Skin damage, circulatory Discharge from semi-conductor

system problems, increased manufacturing, petroleum refining,

risk of cancer wood preservatives, animal feed

additives, herbicides, erosion of

natural deposits

Asbestos 7 million 7 Increased risk of developing Decay of asbestos cement in water

(fiber >10 μm) fibers/L benign intestinal polyps mains, erosion of natural deposits

Barium 2 2 Increase in blood pressure Discharge of drilling wastes,

 discharge from metal refineries,

erosion of natural deposits

Beryllium .004 .004 Intestinal lesions Discharge from metal refineries and

coal-burning factories; discharge from

electrical, aerospace, and defense

industries

Cadmium .005 .005 Kidney damage Corrosion of galvanized pipes,

erosion of natural deposits, discharge

from metal refineries, runoff from

waste batteries and paints

Chromium (total) .1 .1 Some people who use water Discharge from steel and pulp mills,

containing chromium well in erosion of natural deposits

excess of the MCL over many

years could experience allergic

dermatitis.

Copper 1.3 Action levelc Short-term exposure— Corrosion of household plumbing

= 1.3, TT gastrointestinal distress, systems, erosion of natural deposits,

long-term exposure— leaching from wood preservatives

liver or kidney damage

Cyanide (as .2 .2 Nerve damage or thyroid Discharge from steel and metal

free cyanide) problems factories, discharge from plastic and

fertilizer factories

Fluoride 4.0 4.0 Bone disease (pain and Water additive which promotes strong

tenderness of the bones); teeth, erosion of natural deposits,

children may get mottled teeth. discharge from fertilizer and

aluminum factories

Lead Zerod Action levelc Infants and children— Corrosion of household plumbing

= 0.015, TT delays in physical or mental systems, erosion of natural deposits

development; adults—kidney

problems, high blood pressure

Inorganic mercury .002 .002 Kidney damage Erosion of natural deposits, discharge

from refineries and factories, runoff

from landfills and cropland .

Nitrate (measured 10 10 Blue-baby syndrome in infants Runoff from fertilizer use; leaching

as nitrogen) under 6 mo—life threatening from septic tanks, sewage; erosion of

without immediate medical natural deposits

attention

continued
10

Drinking Water Quality

Table 2.3—National primary drinking water regulationsa (States are expected to focus attention on risks

related to the contaminants listed in their source water assessments.) (continued)

MCL Potential health effects Sources of contaminant

Contaminants MCLG or TT from ingestion of water in drinking water

- - - Milligrams per liter - - -

Inorganic chemicals

 (cont.)

Nitrite (measured 1 1 Blue-baby syndrome in infants Runoff from fertilizer use; leaching

as nitrogen) under 6 mo—life threatening from septic tanks, sewage; erosion of

without immediate medical natural deposits

attention

Selenium 0.05 0.05 Hair or fingernail loss, Discharge from petroleum refineries,

numbness in fingers or toes, erosion of natural deposits, discharge

circulatory problems from mines

Thallium .0005 .002 Hair loss; changes in blood; Leaching from ore-processing sites;

kidney, intestine, or liver discharge from electronics, glass, and

problems pharmaceutical companies

Organic chemicals

Acrylamide Zerod TT Nervous system or blood Added to water during

problems, increased risk of sewage and wastewater treatment

cancer

Alachlor Zerod .002 Eye, liver, kidney, or spleen Runoff from herbicide used on row

problems; anemia; increased crops

risk of cancer

Atrazine .003 .003 Cardiovascular system Runoff from herbicide used on row

problems, reproductive crops

difficulties

Benzene Zerod .005 Anemia, decrease in blood Discharge from factories, leaching

platelets, increased risk of from gas storage tanks and landfills

cancer

Benzo(a)pyrene Zerod .0002 Reproductive difficulties, Leaching from linings of water

increased risk of cancer storage tanks and distribution lines

Carbofuran .04 .04 Problems with blood or nervous Leaching of soil fumigant used on

system, reproductive difficulties rice and alfalfa

Carbon Zerod .005 Liver problems, increased risk Discharge from chemical plants and

 tetrachloride of cancer other industrial activities

Chlordane Zerod .002 Liver or nervous system Residue of banned termiticide

problems, increased risk of

cancer

Chlorobenzene .1 .1 Liver or kidney problems Discharge from chemical and

agricultural chemical factories

2, 4-D .07 .07 Kidney, liver, or adrenal gland Runoff from herbicide used on row

problems crops

Dalapon .2 .2 Minor kidney changes Runoff from herbicide used on rightsof-way

1, 2-Dibromo-3- Zerod .0002 Reproductive difficulties, Runoff and leaching from soil

 chloropropane increased risk of cancer fumigant used on soybeans, cotton,

 (DBCP) pineapples, and orchards

o-Dichlorobenzene .6 .6 Liver, kidney, or circulatory Discharge from industrial chemical

system problems factories

p-Dichlorobenzene .075 .075 Anemia; liver, kidney, or spleen Discharge from industrial chemical

damage; changes in blood factories

1, 2-Dichloroethane Zerod .005 Increased risk of cancer Discharge from industrial chemical

factories

continued
11

Table 2.3—National primary drinking water regulationsa (States are expected to focus attention on risks

related to the contaminants listed in their source water assessments.) (continued)

MCL Potential health effects Sources of contaminant

Contaminants MCLG or TT from ingestion of water in drinking water

- - - Milligrams per liter - - -

Organic chemicals

(cont.)

1-1- 0.007 0.007 Liver problems Discharge from industrial chemical

Dichloroethylene factories

cis-1, 2- .07 .07 Liver problems Discharge from industrial chemical

Dichloroethylene factories

trans-1, 2- .1 .1 Liver problems Discharge from industrial chemical

Dichloroethylene factories

Dichloromethane Zerod .005 Liver problems, increased Discharge from pharmaceutical

risk of cancer and chemical factories

1-2- Zerod .005 Increased risk of cancer Discharge from industrial chemical

Dichloropropane factories

Di (2-ethylhexyl) .4 .4 General toxic effects or Leaching from PVC plumbing

adipate reproductive difficulties systems, discharge from chemical

factories

Di (2-ethylhexyl) Zerod .006 Reproductive difficulties, Discharge from rubber and chemical

phthalate liver problems, increased risk factories

of cancer

Dinoseb.007 .007 Reproductive difficulties Runoff from herbicide used on

soybeans and vegetables

Dioxin Zerod .00000003 Reproductive difficulties, Emissions from waste incineration

(2,3,7,8-TCDD) increased risk of cancer and other combustion, discharge

from chemical factories

Diquat .02 .02 Cataracts Runoff from herbicide use

Endothall .1 .1 Stomach and intestinal Runoff from herbicide use

problems

Endrin .002 .002 Nervous system effects Residue of banned insecticide

Epichlorohydrin Zerod TT Stomach problems, Discharge from industrial chemical

reproductive difficulties, factories, added to water during

increased risk of cancer treatment process

Ethylbenzene .7 .7 Liver or kidney problems Discharge from petroleum refineries

Ethylene dibromide Zerod .00005 Stomach problems, Discharge from petroleum refineries

reproductive difficulties,

increased risk of cancer

Glyphosate .7 .7 Kidney problems, Runoff from herbicide use

reproductive difficulties

Heptachlor Zerod .0004 Liver damage, increased Residue of banned termiticide

risk of cancer

Heptachlorepoxide Zerod .0002 Liver damage, increased Breakdown of hepatachlor

risk of cancer

Hexachlorobenzene Zerod .001 Liver or kidney problems, Discharge from metal refineries and

reproductive difficulties, agricultural chemical factories

increased risk of cancer

Hexachlorocyclo- .05 .05 Kidney or stomach problems Discharge from chemical factories

pentadiene

Lindane .0002 .0002 Liver or kidney problems Runoff and leaching from insecticide

used on cattle, lumber, gardens

continued

Chapter 2
12

Drinking Water Quality

Table 2.3—National primary drinking water regulationsa (States are expected to focus attention on risks

related to the contaminants listed in their source water assessments.) (continued)

MCL Potential health effects Sources of contaminant

Contaminants MCLG or TT from ingestion of water in drinking water

- - - Milligrams per liter - - -

Organic chemicals

 (cont.)

Methoxychlor 0.04 0.04 Reproductive difficulties Runoff and leaching from insecticide

used on fruits, vegetables, alfalfa,

livestock

Oxamyl (Vydate) .2 .2 Slight nervous system effects Runoff and leaching from insecticide

used on apples, potatoes, and

tomatoes

Polychlorinated Zerod .0005 Skin changes, thymus gland Runoff from landfills, discharge of

biphenyls (PCB’s) problems, immune deficiencies, waste chemicals

reproductive or nervous system

difficulties, increased risk of

cancer

Pentachlorophenol Zerod .001 Liver or kidney problems, Discharge from wood-preserving

increased risk of cancer factories

Picloram .5 .5 Liver problems Herbicide runoff

Simazine .004 .004 Problems with blood Herbicide runoff

Styrene .1 .1 Liver, kidney, and circulatory Discharge from rubber and plastic

problems factories, leaching from landfills

Tetrachloroethylene Zerod .005 Liver problems, increased Leaching from PVC pipes, discharge

risk of cancer from factories and dry cleaners

Toluene 1 1 Nervous system, kidney, or Discharge from petroleum factories

liver problems

Total Noneb .10 Liver, kidney, or central nervous By-product of drinking water

trihalomethanes system problems; increased disinfection

(TTHM’s) risk of cancer

Toxaphene Zerod .003 Kidney, liver, or thyroid Runoff and leaching from insecticide

problems; increased risk of used on cotton and cattle

cancer

2,4,5-TP (Silvex) .05 .05 Liver problems Residue of banned herbicide

1,2,4- .07 .07 Changes in adrenal glands Discharge from textile finishing

Trichlorobenzene factories

1,1,1- .20 .2 Liver, nervous system, or Discharge from metal degreasing

Trichloroethane circulatory problems sites and other factories

1,1,2- .003 .005 Liver, kidney, or immune Discharge from industrial chemical

Trichloroethane system problems factories

Trichloroethylene Zerod .005 Liver problems, increased Discharge from petroleum refineries

risk of cancer

Vinyl chloride Zerod .002 Increased risk of cancer Leaching from PVC pipes, discharge

from plastic factories

Xylenes (total) 10 10 Nervous system damage Discharge from petroleum factories,

discharge from chemical factories

continued
13

Chapter 2

Table 2.3—National primary drinking water regulationsa (States are expected to focus attention on risks

related to the contaminants listed in their source water assessments.) (continued)

MCL Potential health effects Sources of contaminant

Contaminants MCLG or TT from ingestion of water in drinking water

- - - Milligrams per liter - - -

Radionuclides

Beta particles and Noneb 4 millirems Increased risk of cancer Decay of natural and man-made

photon emitters per yr deposits

Gross alpha particle Noneb 15 pCi/L Increased risk of cancer Erosion of natural deposits

activity

Radium 226 and Noneb 5 pCi/L Increased risk of cancer Erosion of natural deposits

radium 228

(combined)

Microorganisms

Giardia lamblia Zerod TT Giardiasis—a gastroenteric Human and animal fecal waste

disease

Heterotrophic plate NA TT No health effects but can NA

count indicate how effective

treatment is at controlling

microorganisms.

Legionella Zerod TT Legionnaire’s Disease—a Found naturally in water, multiplies

form of pneumonia in heating systems

Total coliforms Zerod 5.0% Used as an indicator that other Human and animal fecal waste

(including fecal potentially harmful bacteria

coliform and may be present

E. coli)

Turbidity NA TT Turbidity has no health effects Soil runoff, growth of algae

but can interfere with disinfection

and provide a medium for

microbial growth. It may indicate

the presence of microbes.

Viruses (enteric) Zerod TT Gastroenteric disease Human and animal fecal waste

MCL = maximum contaminant level or the maximum permissible level of a contaminant in drinking water delivered to any user;

MCLG = maximum contaminant level goal; NA = not available; pCi = picocuries; PVC = polyvinyl chloride; TT = treatment technique. a Water-quality regulations are subject to change. For the latest regulations, visit the Web site: http://www.epa.gov/OGWDW/wot/appa.html. b MCLG has not been defined.

c

The units vary with the contaminant and are defined by the U.S. Environmental Protection Agency.

d MCLG is 0.0.

Source: U.S. EPA 1999b.
14

Dissolved Gases

The most abundant dissolved gases in water are nitrogen

(N2), oxygen (O2), carbon dioxide (CO2), methane (CH4),

hydrogen sulfide (H2S), and nitrous oxide (N2O). The first

three are abundant in the Earth’s atmosphere. The second

three are typically products of biogeochemical processes

that occur in nonaerated, low oxygen environments. The

solubility of most gases increases with decreasing water

temperature and decreases with increasing concentrations

of chlorides or other salts.

The concentration of dissolved oxygen (DO) is essential to

aquatic life and can effect the water’s color, taste, odor, and

chemistry. Unpolluted surface waters are generally saturated

with DO because of reaeration and the production of oxygen

during photosynthesis by submerged aquatic plants. Ground

water systems tend toward oxygen depletion and reducing

conditions because the oxygen consumed during

hydrochemical and biochemical reactions is not replenished

by the atmosphere. Polluted surface waters tend to have

lower DO concentrations because of oxygen consumption

during the decomposition of organic matter.

The concentrations of DO strongly influence the solubility

and stability of elements that readily gain or lose electrons

including iron (Fe+3), manganese (Mn+3), nitrogen, sulfur

(S), and arsenic (As+3). When dissolved iron and manganese

are exposed to air, they form insoluble precipitates that

make water turbid, cause stains in laundry, and impart a

bitter taste (Cox 1964). In water with little or no oxygen,

iron minerals are reduced, and adsorbed phosphorus and

other elements can be released into the water. The solubility

of most arsenic and arsenic-sulfur compounds depends on

the presence of DO and can have concentrations in water

above the primary standard of 0.05 milligrams (mg) per liter

(L) (Freeze and Cherry 1979).

Organic Compounds

Organic compounds have carbon and usually hydrogen and

oxygen as the main components in their structural framework. They are typically nonpolar, have relatively low

solubility, and are degraded by microorganisms, hydrolysis,

oxidation-reduction, and volatilization. In natural waters,

they are transported as dissolved phases and attached to

particulates.

Commonly occurring natural organic compounds include

plant and animal tissue and the products of their decomposition. Synthetic organics found in water include petroleum

products, pesticides, and herbicides (table 2.3). Most

synthetic toxic organic compounds originate from coal

mining, petroleum refining, and manufacture of textile,

wood pulp, and pesticides (table 2.1). In the environment,

they are usually associated with roadways and industrial,

urban, and agricultural land uses. Disinfecting some

organic-rich waters with chlorine may also result in the

formation of carcinogenic organic compounds such as

trihalomethanes (Martin and others 1993, see chapter 5).

Highly soluble, potentially carcinogenic organic compounds

from gasoline spills and emissions are also found in water

supplies and can make water distasteful and undrinkable

(see chapter 6).

Trace Metals and Nonmetals

Primary and secondary water-quality standards have been

developed for common trace metals and nontrace elements

(tables 2.3, 2.4). Most of these elements occur in natural,

uncontaminated waters in concentrations below 1 mg per

liter. Metals have relatively low solubilities. Solubilities are

usually lowest at neutral acidity and increase with increasing

acidity and increasing alkalinity. A characteristic feature of

metals is their tendency to form hydrolyzed species and

Drinking Water Quality

Table 2.4—National secondary drinking water

regulations, which are nonenforceable guidelines for

contaminants that may cause cosmetic effects (e.g., skin

or tooth discoloration) or aesthetic effects (e.g., taste,

odor, or color) in drinking water

Contaminant Unit Secondary standard

Aluminum mg/L 0.05–0.2

Chloride mg/L 250

Color Color units 15

Copper mg/L 1.0

Fluoride mg/L 2.0

Foaming agents mg/L .5

Iron mg/L .3

Manganese mg/L .05

Odor Threshold 3

odor number

pH 6.5 –8.5

Silver mg/L .10

Sulfate mg/L 250

Total dissolved mg/L 500

solids

Zinc mg/L 5

Source: U.S. EPA 1999c.
15

Table 2.5—Common chemical processes involved as water interacts with its environment

Process Description

Acid-base reactions Acid-base reactions are a common type of chemical reaction in aqueous environments

that are important in the leaching and transport of cations. They are also important in

certain water treatment processes and in the corrosion of water distribution systems.

Acids are hydrogen-containing substances that supply protons in water, typically by

liberating hydrogen ions. Bases are proton acceptors and are typically substances that

contain hydroxide ions (OH–) or hydroxyl groups, which dissociate in water. Acidity is

usually measured using the logarithmic pH scale, which is defined as the concentration

of hydrogen ions in water in moles per liter (see glossary of terms). Acidic soil or waters

can have increased concentrations of metals and decreased phosphate availability and

nitrification rates. The dissolution of carbon dioxide in water to form carbonic acid

(H2CO3) is the most common acid-producing reaction in natural waters.

Adsorption-desorption Adsorption-desorption is the exchange of chemicals from solution and the surfaces of

charged particles by chemical or physical bonding. When the adsorption bonds are

chemical, they are relatively irreversible. If they are physical van der Waals type forces,

they are easily broken and reversible. Particle type (organic or inorganic), particle size

(clay, sand, etc.), and the presence of organic and inorganic coatings can have large

effects on the amount of adsorption and desorption of organic waste, pesticides,

ammonia, and phosphorus as they are transported by water through soils. In general,

adsorption tends to increase with increases in the content of both clay and organic

matter. The removal of contaminants in water by adsorption and subsequent settling of

sediments is an important process in lakes, rivers, and water treatment plants.

Volatilization Volatilization is the loss of a chemical from the soil-water system by vaporization into

the atmosphere. The rate of volatilization depends on the concentration gradient above

the volatilization surface and typically increases with temperature and the removal of

vaporized chemicals away from the surface by wind or heat. This is a particularly

important process in fires and after the application of pesticides or nutrients.

Reaeration Reaeration is the transfer of gases, typically oxygen, from the atmosphere into water.

The rate of reaeration increases with turbulence, exposed surface area, and the solubility

and diffusivity of gas, both of which are temperature dependent. Oxygen is the most

common dissolved gas in water and is essential for aquatic life and the decomposition of

natural and synthetic organic matter.

Oxidation-reduction Oxidation is the loss of electrons and reduction is the gain of electrons. The redox

potential is used to express the tendency to exchange electrons and is measured as the

voltage required to prevent the acceptance of electrons on a standard electrode. Oxic

environments are considered to have high redox potentials because O2 is available as an

electron acceptor. In order to reduce inorganic constituents, some other constituents

must be oxidized, typically organic matter in reactions that are catalyzed by bacteria or

isolated enzymes.

Decomposition- Decomposition is a general term that refers to the breakdown of organic matter.

mineralization- Mineralization specifically refers to decomposition processes that release carbon

immobilization as CO2 and nutrients in inorganic forms. This breakdown usually involves soil microbes

and is caused by some combination of photolysis, hydrolysis, oxidation-reduction, and

enzyme actions. Immobilization is the accumulation of N, P, and other nutrients in soil

microbes.

Chapter 2
16

Table 2.6—Summary of the chemical behavior of important water contaminants

Biochemical Chemical Physiochemical

transformationsa reactions retardationb

Contaminant Aerobic Anaerobic Acid Alkaline Acid Alkaline

Metals

Aluminium (Al+3) MP P D P D

Cadmium (Cd+2) P P P DP M

Chromium (Cr+3) PP MPD P

Copper (Cu+1) P P M DD M

Iron (Fe+3) DP P DM D

Lead (Pb+4) P P M DM D

Manganese (Mn+3) MM P D M D

Mercury (Hg+1) MP M D M D

Silver (Ag+1) P P M DM D

Zinc (Zn+2) P P P DP D

Inorganic nonmetals

Ammonium (NH4

+1)D P P P P D

Nitrate (NO3

-1) PD P PP P

Sodium (Na) P P P P P M

Sulphate (SO4

-2) P D P MP P

Fluoride (F-1) P P M MD P

Chloride (Cl-1) PP P PP P

Arsenic (As) P P M P M D

Selenium (Se) P P D M D P

Cyanide P P P P D P

Organic compounds

Aliphatic D P P P D D

Hydrocarbons

Phenols D M P P M M

Benzene D P P P D D

Toluene D P P P D D

Polynuclear

aromatics M P P P M M

Halogenated organics

Tri and tetra

chloroethylene P M P P M M

Carbon tetrachloride P M P P M M

Chloroform P M P P P P

Methylene chloride M M P P P P

Chlorobenzene D P P P D D

Chlorphenols D M P P P M

Fecal organisms

Fecal coliform P P P P M P

Pathogenic bacteria P P P P M P

Pathogenic virus P P D M M D

D = reactions do occur; M = reactions may occur; P = reactions probably occur.

a

Biochemical transformations involve biological organisms, usually microbes.

b

Physiochemical retardation involves physical and chemical bonds that are usually to mineral surfaces.

Source: Adapted and updated from Foster and Gomes 1989.

Drinking Water Quality
17

inorganic and organic complexes. These complexes typically absorb to suspended particulates or form insoluble

precipitates. Therefore, the transport of metals across the

landscape is often related to acidity, the presence of organic

compounds, and the transport of sediment (table 2.6).

While trace metals and nonmetals occur naturally, their

concentrations can be greatly increased over background

levels by mining activities, waste dumps, acidic runoff,

tanneries, and other industries. Some metals, such as copper

and cadmium, are associated with automobiles and are

concentrated on streets, parking lots, and industrial areas

(Bannerman and others 1993). Major sources of lead include

urban soil, lead-based paint, and some hair-coloring

cosmetics (Mielke 1999).

Fluorine (F) is a trace nonmetal that occurs as fluoride and

is undersaturated in nearly all natural water. Because it can

have beneficial effects on dental health, fluorine is added to

some municipal water supplies. Arsenic is a soluble trace

nonmetal that can be naturally present in water from areas of

recent volcanism. It is widely used in pigments, insecticides,

herbicides, and metal alloys (Freeze and Cherry 1979).

Selenium (Se) is a toxic nonmetallic element that has

geochemical properties similar to sulfur. It can occur in

appreciable concentrations in coal, uranium ore, certain

shales, and discharges from petroleum refineries and mines.

Like sulfur, it forms strong chemical bonds on the surface of

minerals and can be reduced by anaerobic bacteria

(Schlesinger 1997).

Nitrogen

Nitrogen, a major nutrient for vegetation, plays a dominant

role in many biochemical reactions. However, in certain

chemical forms, it can adversely affect humans, ecosystems,

and water supplies. Since preindustrial times, fertilizer

production and other human activities have more than

doubled the global input of nitrogen to terrestrial ecosystems

(Kinzig and Socolow 1994, Vitousek and others 1997). This

increase has made nitrogen the most common water

pollutant in the United States. In the Northeastern United

States alone, anthropogenic activities have apparently

increased the nitrate concentrations in major rivers threefold

to tenfold since the early 1900’s (Matson and others 1997).

Anthropogenic alteration of nitrogen cycles has also affected

forest and aquatic productivity and increased acid rain,

photochemical smog, and greenhouse gases (Fenn and

others 1998, Vitousek and others 1997).

Certain nitrogen compounds can have toxic effects at

relatively low concentrations. Methaemoglobinemia (bluebaby syndrome) in bottle-fed babies and the elderly is a

human health hazard associated with nitrite (NO2

-1) in

drinking water (table 2.2). Nitrate in water can also present

similar health hazards as can nitrate in many foods (GEMS

1991). Bacteria residing in vertebrate digestive tracts can

convert the relatively benign nitrate into the toxic nitrite

(Kinzig and Socolow 1994). Ammonia dissolved in drinking

water is not toxic to humans but can be toxic to some

aquatic invertebrates and fish depending on the concentration of DO temperature, acidity, and salinity, and the carbon

dioxide-carbonic acid equilibrium of water. Because all

forms of inorganic nitrogen are nutrients to green plants,

excessive concentrations in water can lead to algal blooms,

excessive growth of submerged aquatic plants, and eutrophication, particularly in coastal and marine ecosystems.

The global nitrogen cycle consists of three major reservoirs:

(1) the atmosphere, (2) the hydrosphere, and (3) the biosphere (fig. 2.1). The flow between these reservoirs occurs

in many forms and pathways (fig. 2.2). Inorganic nitrogen

can be transported in water as dissolved nitrous oxide or

nitrogen gas, ammonia, and cations or as anions of nitrite or

nitrate. The concentrations of these compounds are low in

most unpolluted freshwater and high in waters contaminated

by organic wastes, sewage, or fertilizers. Worldwide,

pristine rivers have average concentrations of ammonia and

nitrate of 0.015 mg per liter and 0.1 mg per liter, respectively (GEMS 1991). Nitrate concentrations >1 mg per liter

generally indicate anthropogenic inputs. The lowest concentrations are generally found in deep ground water and

surface waters draining pristine wildlands (GEMS 1991,

Spahr and Wynn 1997). The highest levels are associated

with surface runoff and ground water from fertilized

agricultural and urban areas. In undisturbed watersheds,

annual yields increase with annual runoff, and yields from

savanna and rangeland are less than from forest (Lewis and

others 1999).

Organic nitrogen is converted to inorganic nitrogen in a

process called mineralization in the following oxidation

sequence: organic nitrogen and ammonium to nitrite to

nitrate. In water that is strongly oxidized, nitrate is the stable

phase and is very mobile. As redox potential declines,

nitrate is reduced or denitrified to nitrous oxide or nitrogen

gas. Because of the potential adverse ecosystem and health

effects associated with nitrites and nitrates, denitrification is

desirable for water quality. Generally, the amount of net

mineralization is directly related to the total content of

organic nitrogen and carbon (Schlesinger 1997, Vitousek

and Melillo 1979). Nitrification tends to be lower in soil

with low acidity, low soil oxygen, low soil moisture, and

low temperature, and high litter carbon to nitrogen ratios. At

the watershed scale, rates of denitrification vary with

landscape positions (Jordan and others 1993, Peterjohn and

Chapter 2
18

Figure 2.1—Global nitrogen cycle. Annual fluxes in units of 1012 grams per year.

Correll 1984). In general, relatively high denitrification rates

are found in riparian forests and at the base of slopes where

water, carbon, nitrogen, and phosphorus are readily

available.

Because nitrogen is essential to the growth of plants,

seasonal differences in plant uptake can cause measurable

variations in the concentration of nitrogen in soil and

surface water. In general, the lowest nitrogen levels in

surface or ground water occur during the early growing

season when plant uptake is greatest (Boyd 1996). Maximum nitrogen concentrations typically occur in the winter

when plant uptake is reduced, and the dissolved fraction is

concentrated in unfrozen water. However, seasonal trends

can be reversed or diminished in areas with large

anthropogenic inputs.

Phosphorus

The presence of phosphorus in drinking water is not

considered a human health hazard, and no drinking waterquality standards are established for phosphorus. Nevertheless, phosphorus can affect the water’s color and odor and

indicate the presence of other organic pollution. Furthermore, because phosphorus can accelerate the growth of

algae and aquatic vegetation, it contributes to the eutrophication and associated deterioration of municipal water

supplies. Whereas excess nitrogen is responsible for most of

Drinking Water Quality

Land

plants

100

Human

activities

Atmosphere

Denitrification

110

Biological

fixation

15

The Global Nitrogen Cycle

Internal

cycling

Ground water

Soil organic - N

Internal

cycling

Oceans

Biological

fixation

140

1200

Denitrification

 ≤200

Permanent

burial

10

8000

River

flow

36

30

Fixation in

lightning

<3
19

Chapter 2

Figure 2.2—Sources and pathways of nitrogen in the subsurface environment.

Organic - N

proteins

N2 N2

Denitrification

Decomposition

nitrification

Adsorption

Proteins

water

Precipitation

Mineral

Sewage fertilizer Plant residue,

compost

Plant

proteins

Nitrification

Nitrogen

fixation

Denitrification

Denitrification in reducing zones

Decomposition

Nitrification

Denitrification

 Adsorption

N2

{ }

NH3

NO-3

NH3

NO-3

NO-3

Organic - N NH3

NH3

N2 NH3

NH+

4

NO-3

NO-3

NH+

4

NO-3 NO-3

NH+

4

 N2

NH+

4

Leaching

N2

(aq) N20

NO-3 NO-3 NO-3

Ground
20

the coastal and marine eutrophication, agricultural sources

of phosphorus dominate the eutrophication processes in

many freshwater aquatic systems (Matson and others 1997).

Nearly all the phosphorus in terrestrial ecosystems is

originally derived from the weathering of minerals (fig. 2.3).

The most common phosphorus-rich mineral is apatite, a

calcium orthophosphate that is present in some igneous

rocks and marine sediments. In natural freshwater, phosphorus exists in both dissolved and particulate fractions.

Dissolved phases typically originate from excretions by

organisms, whereas particulate fractions can have organic or

inorganic origins. In streams, a large fraction of phosphorus

is adsorbed on and transported with organic and inorganic

particulates. In lakes, a large proportion of the phosphorus in

oxygen-rich surface waters is held in plankton biomass

(Schlesinger 1997). In deeper, anoxic lakes, phosphorus is

adsorbed to sediments and particulates but can be released

during the reduction of iron compounds. Unlike nitrogen,

carbon, and hydrogen, phosphorus does not have a significant gaseous component.

Chemical Evolution of Water

As water moves across the landscape, it interacts with the

surfaces it contacts and chemically evolves toward the

composition of seawater [for detailed explanations see

Stumm and Morgan (1970) and Freeze and Cherry (1979)].

In general, the evolution of deep ground water typically

involves increases in dissolved solids and decreases in DO,

organic waste, pesticides, phosphorus, and nitrogen. In

contrast, the concentrations of organic waste, pesticides,

phosphorus, and nitrogen increase as surface water travels

across the landscape and interacts with both natural and

anthropogenic systems.

Fresh, young water that has had little contact with its

surroundings is generally low in total dissolved solids and

rich in bicarbonate anions derived from soil carbon dioxide

and the dissolution of carbonate minerals. Sulfate anions

tend to dominate in intermediate age ground water while

chloride anions dominate in older, deep ground water that

has traveled long distances. These sulfate and chloride

anions are derived from the dissolution of soluble sedimentary minerals. Because these minerals are present only in

small amounts in most rocks, water usually has to travel

considerable distances before it is dominated by either

sulfate or chloride anions.

The DO content and redox potential tend to decrease as

water travels across the landscape. Rain and snow are

exposed to atmospheric oxygen and have relatively high DO

and redox potentials. As water passes through organic-rich

forest litter and soil, the DO is removed, redox potential

declines, and large amounts of organic acids are generated.

Nutrient immobilization predominates in the upper layers of

fresh litter, while mineralization of nitrogen, phosphorus,

and sulfur is usually greatest in the upper mineral soil. As

water travels through the subsurface, all the DO is consumed by bacterially catalyzed reactions that oxidize

organic matter. Eventually the aerobic bacteria involved in

these reactions can no longer thrive, and anaerobic conditions prevail. Then ammonia, manganese, ferrous iron, and

sulfate become oxidizing agents.

Cation concentrations in water vary considerably in space

and time and do not follow well-defined, theoretically based

sequences like anions or redox potentials. Nevertheless,

cations enter the aquatic system from the weathering of

minerals and the breakdown of organic materials. Their

concentrations typically increase with travel distance in both

surface and ground water. The most abundant cations in

water supplies are calcium and magnesium, which can be

removed by chemical treatments to prevent scaling of pipes

and to reduce the amount of soap needed for washing.

Physical Properties

The physical characteristics of concern in drinking water are

temperature, color, turbidity, sediments, taste, and odor.

Temperature

Because of its hydrogen bonds and molecular structure,

water has an unusual trait—the density of its solid phase

(ice) is lower than that of its liquid phase (water). Because

of this trait, ice floats, and pipes and plant tissues rupture

when the water within them freezes and expands.

The rates of chemical and metabolic reactions, viscosity and

solubility, gas-diffusion rates, and the settling velocity of

particles depend on temperature. Metabolism, reproduction,

and other physiological processes of aquatic organisms

are controlled by heat-sensitive proteins and enzymes

(Ward 1985). A 10 °C increase in temperature will roughly

double the metabolic rate of cold-blooded organisms and

many chemical reactions. A permanent 5 °C change in

temperature can significantly alter the structure and composition of an aquatic population (MacDonald and others

1991, Nathanson 1986). Temperature increases also decrease DO concentrations but can increase the oxidation rate

and efficiency of certain biological, wastewater treatment

systems.

Drinking Water Quality
21

Figure 2.3—Global phosphorus cycle. Annual fluxes in units of 1012 grams per year.

Chapter 2

The temperature of water naturally varies with time of day,

season, and the type of water body. Changes in surface

water temperatures reflect seasonal changes in net radiation,

daily changes in air temperature, and local variations in

incoming radiation. Temperature variations in ground water

are less than in surface water. Except in the winter, surface

water is usually warmer than ground water, and most

anthropogenic activities increase water temperatures.

Removal of vegetative canopies over streams influences

water temperatures by affecting energy inputs, evaporative

cooling, and the way water flows across the landscape. The

cooling rate for surface water depends on heat transfer to the

atmosphere.

Seasonal and spatial variations in the temperature in water

supply reservoirs can have large effects on the quality of

raw municipal water (Cox 1964). Water in deep reservoirs is

commonly divided into three zones: the upper circulating

zone, the middle transition zone, and the deepest zone of

stagnation. Water in the upper surface zone is aerated and

mixed by wind action and typically has abundant DO. In

contrast, the deepest, stagnant water contains little or no DO

because it has been removed during the oxidation of organic

matter. The breakdown of organic matter also makes deep

water acidic and rich in carbonic acid. Consequently,

stagnate, deep water has the chemical conditions necessary

to dissolve iron, manganese, sulfur, and other taste- and

odor-producing substances. To avoid the objectionable taste

and odor of the deep water, municipal water is usually

drawn from the surface of the reservoir. However, when the

temperature of the surface water falls rapidly, it can become

denser than the bottom water, causing the entire water

column of the reservoir to mix or “turn over.” During these

mixing events, the DO content of the entire lake can

decrease, causing massive fish kills and foul smelling and

poor tasting water. Similar mixing can occur in stratified

lakes or estuaries during periods of intense runoff.

Color and Turbidity

Pure water is colorless in thin layers and bluish green in

thick layers. Particulates and insoluble compounds typically

add color and reduce transparency. Consequently, the

presence of light-dependent aquatic organisms can affect

esthetic appeal and taste of water as well as the effectiveness

of certain wastewater treatment processes.

 60

Soils

200,000

Internal

cycling

Mining

Mineable

rock

10,000

Riverflow

Reactive

Bound

Dust transport

The Global Phosphorus Cycle

Internal

cycling

90,000

Sediments

4.0 x 109

12

1.0

 1000

2

19

2

Land plants

3,000
22

Turbidity is an optical property related to the scattering of

light and clarity. It is typically controlled by the presence of

suspended particles or organic compounds. Turbidity itself

is not injurious to human health. Approximately 50 percent

of the total incident light is scattered or transformed into

heat within the first meter of water. As turbidity increases, it

reduces the depth of sunlight penetration, thereby altering

water temperature and stratification, the photosynthesis of

aquatic organisms, the DO content of the water body, and

the cost of water treatment. In addition, turbid water can

contain particulate of soil or fecal matter that harbors

microorganisms and/or carries absorbed contaminants. The

removal of particulates by gravity or by addition of chemicals is typically the first step in treating water for human

consumption. The sedimentation of particles and the

bleaching action of sunlight during reservoir storage can

reduce both the color and turbidity of water (Cox 1964).

Sediment

Sediment is a major water-quality concern because of its

ability to transport harmful substances and its impacts on the

cost of water treatment and the maintenance of water

distribution systems. While sediment is derived during the

natural weathering and sculpturing of the landscape,

accelerated levels of erosion and sedimentation are associated with many anthropogenic activities (table 2.1).

The general term sediment includes both organic and

inorganic particles that are derived from the physical and

chemical weathering of the landscape. Individual particles

are eroded, transported, and deposited. Erosion can be either

physical or chemical. Transport can be by wind, gravity, or

water. In water and air, particles can be transported in

suspension (suspended load) or along the substrate (bed

load). Sediment load is the total quantity of sediment that is

transported through a cross-section of a stream during a

specific time period. The actual amount of sediment

transported at any place or time depends on the supply of

sediment and the transport capacity of the stream. Sediment

is usually measured as mass per unit area (tons per acre per

year or metric tonnes per hectare per year), concentration

(parts per million or milligrams per liter), or lowering of the

landscape (inches per 1,000 years or millimeters per 1,000

years). In general, high sediment loads increase water

treatment costs and reduce the storage volume and life span

of water storage facilities.

Biological Properties

Aquatic organisms are usually grouped into those that

(1) obtain the carbon they need for biosynthesis from carbon

dioxide (autotrophs) and (2) use existing organic compounds as their carbon source (heterotrophs). Generally,

autotrophs increase DO concentrations in water through

photosynthesis, while heterotrophs are responsible for

breakdown and recycling of dead organic materials and

decreased DO concentrations.

Most microbial contaminants in water are caused by

heterotrophs that are transmitted to a water system via

human and animal fecal matter (U.S. EPA 1999a). Most

waterborne pathogenic microorganisms are bacteria or

viruses that survive in sewage and septic leachate (table

2.7). Bacterial pathogens are generated by both animal and

human sources, while viral pathogens are usually only

generated by human sources. Viruses that infect animals

normally do not cause illness in humans. However, animal

sources for some viruses that effect humans are suspected,

particularly viruses that infect the respiratory system like the

sin nombre virus, hantavirus, influenza virus, and Ebola

virus.

Common bacterial diseases spread by aquatic microorganisms include Legionnaire’s disease, cholera, typhoid, and

gastroenteritis. Waterborne viral diseases include polio,

hepatitis, and forms of gastroenteritis. Waterborne parasitic

diseases include amoebic dysentery, flukes, and giardiasis.

Giardia spp. and Cryptosporidium spp. are parasitic

protozoans that are transferred between animals and humans

via the fecal-oral route and are significant sources of

gastrointestinal illness. They are common in surface water in

back-country areas, including in many national forests and

parks. These back-country areas, which provide animal

habitat, experience low human use (Monzingo and Stevens

1986) (see chapter 15). Unfortunately, some parasitic

protozoans are not removed in most water treatment plants

because they are small enough to pass filtration systems and

are very resistant to disinfectants.

The analytical procedures for detecting waterborne viral

diseases are costly and time consuming. Therefore most

drinking and recreational waters are routinely tested for

microbes that are easier to detect but whose presence is

highly correlated with human health hazards. Coliforms are

the most common type of microbes used in this type of

testing. All coliforms are aerobic and facultative anaerobic,

gram-negative, nonspore-forming, rod-shaped bacteria that

ferment lactose. Their presence and abundance in raw water

is used to screen for fresh fecal contamination (Cox 1964).

Their presence in treated water is used to determine treatment plant efficiency and the integrity of the distribution

system.

Many environmental factors can affect the transport of

microbes across the landscape (table 2.8). Relatively

Drinking Water Quality
23

Table 2.7—Common waterborne pathogenic and indicator bacteria and viruses

Waterborne pathogenic bacteria Waterborne pathogenic viruses

Legionella

Mycobacterium avium intracellular (MAC)

Shigella (several strains)

Helicobacter pylori

Vibrio cholerae

Salmonella typhi

S. typhimurum

Yersinia

Campylobacter (several strains)

Escherichia coli (several pathogenic strains)

Waterborne indicator bacteria Waterborne indicator viruses

Total coliform

Fecal coliform

 E. coli (both nonpathogenic and

 pathogenic strains)

Enterococci

Fecal streptococci

Clostridium perfringens (anaerobic spores)

Klebsiella pneumoniae

Aeromonas hydrophila

Source: U.S. EPA 1999a.

Chapter 2

coarse-grained or sandy soils are poor adsorbers of microbes

(Keswick and Gerba 1980, U.S. EPA 1999a). Fine-textured

clay soils or soils with abundant colloidal organic material

are very adsorbent because their negatively charged surfaces

and large surface area per-unit volume increase the number

of potential adsorption sites for microbial contaminants. As

a result, clay soils slow the migration but can enhance the

survival of certain microbes (Bitton and others 1986,

Keswick and Gerba 1980). In contrast, the absorption of

viruses to organic soils or in environments with high

concentrations of dissolved organic matter or organic acids

is relatively poor, probably because of competition for

adsorption sites. The presence of humic and fulvic acids

may reduce virus infectivity.

The acidity and ionic strength of liquids percolating past

adsorbed microbes can influence their sorption and desorption. Moreover, a reduction in the ionic strength of pore

water weakens the virus-soil adsorption forces and increases

their entrainment and concentrations in percolating water

(Bitton and others 1986). Therefore, natural rainwater with

its extremely low ionic strength can mobilize and transport

viruses that have sorbed to the upper layers of the soil.

Fecal contamination of surface and ground water can occur

by several pathways (table 2.1). The concentration of

microbes in surface runoff is generally higher in warmer

months and higher in runoff from grazed rather than

ungrazed land (Edwards and others 1997). Lawns and

residential streets are important sources of fecal coliforms

from domestic animals (Bannerman and others 1993).

Leaking sewer lines and failed septic systems are also

common sources (U.S. EPA 1999a), and water distribution

systems can harbor bacterial or fecal contamination. This

contamination enters distribution systems when controls fail

or when negative pressure in a leaking pipe allows contaminants to infiltrate.

Storage in reservoirs can increase or decrease the microbial

content of surface water. Sedimentation of particles with

adsorbed microbes and the germicidal action of sunlight can

lower microbial content (Cox 1964). However, these effects

are spatially and seasonally variable and are influenced by

microclimate and the morphology and chemistry of a water

body (James and Havens 1996). Eutrophic conditions that

reduce DO concentrations or produce toxic blue-green algae

blooms may decrease water quality (see Hebgen Lake case

in chapter 5).

Bacteriophage

 Bacteroides phage

 Coliphage

Male-specific coliphage

FRNA phage

FDNA phage

Host Salmonlla WG-49

Host E. coli C-3000

Host E. coli FAMP

Host E. coli 15597

Somatic coliphage

Host E. coli C 13706, C-3000

Host Salmonella WG-49

Enteroviruses

Coxsackieviruses

Echoviruses

Poliovirus

Enterovirus 70 and 71

Hepatitis A virus

Hepatitis E virus

Enteric adenoviruses

Rotavirus

Norwalk virus

Small round structured viruses (SRSV)

Astrovirus

Caliciviruses
24

Drinking Water Quality

Table 2.8—Factors influencing virus transport and fate in the subsurface

Factor Influence on fate of virus Influence on transport

Light Minor factor in virus inactivation, effective only at Unknown

the soil’s surface

Temperature Viruses survive in soil and water longer at lower Unknown

temperatures.

Hydrogeologic A short ground water time of travel indicates that Relatively slow flow reduces the rate

conditions and well viruses may be transported to water supply wells of virus migration while conduit,

pumping rate before dying off or becoming inactivated. High fracture flow, or rapid flow in coarsepumping rates decrease ground water travel times. grained, porous media enhances

transport.

Soil properties; iron- Effects on survival are probably related to the degree High degree of virus retention by the

oxide coatings on of virus adsorption. Iron oxides probably increase clay fraction of soil; iron coatings

soil or aquifer grains inactivation. may be especially efficient in

providing an attractive surface for

virus attachment.

pH Most enteric viruses are stable between a pH range Generally, low pH favors virus

of 3 to 9. Survival may be prolonged at near-neutral adsorption and high pH results in

pH values. virus desorption from particles.

Inorganic ions/salt Some viruses are protected from inactivation by Generally, increasing the concentration

species and certain cations; the reverse is also true. of ionic salts and increasing cation

concentration valencies enhance virus adsorption.

Organic matter Presence of organic matter may protect viruses from Soluble organic matter competes with

inactivation; others have found that it may reversibly viruses for adsorption sites on soil

retard virus infectivity. particles.

Virus type Different virus types vary in their susceptibility to Virus adsorption to soils is probably

inactivation by physical, chemical, and biological related to physicochemical differences

factors. in virus capsid surfaces.

Microbial activity Some viruses are inactivated more readily in the Unknown

presence of certain microorganisms; however,

adsorption to the surface of bacteria can be

protective.

Iron content in May increase virus attachment and inactivation Iron-oxidizing bacteria may form a

shallow soil or biomass layer that filters out viruses.

aerobic aquifers Heavy precipitation events may cause

the ionic strength of the water to

decline and the biofilms to release the

filtered organisms.

Soil moisture Influences inactivation and adsorption to particle Increased saturation promotes

content surfaces; survival may increase in unsaturated desorption of viruses from particle

conditions. surfaces and migration in

ground water.

Source: U.S. EPA 1999a.
25

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Chapter 2
26

Chapter 3

Watershed Processes—Fluxes of Water,

Dissolved Constituents, and Sediment

F.J. Swanson, F.N. Scatena, G.E. Dissmeyer, M.E. Fenn,

E.S. Verry, and J.A. Lynch1

The Integrated Hydrologic System

Ecosystems are energy-processing units that are continually

cycling and being regulated by essential nutrients and water.

Some cycles, like the hydrologic cycle, are global and, thus,

involve transport over great distances. Other cycles occur

locally among biotic elements, forest litter, and soil. In most

forests, large pools of tightly bound, relatively unavailable

nutrients are linked with small pools of available nutrients

that are rapidly cycled through the ecosystem.

The circulation of water through the hydrologic cycle is the

largest movement of a chemical substance at the surface of

the earth (Schlesinger 1997). The hydrologic cycle describes

the constant exchange of water among the land, sea, and

atmosphere. A water budget is the balance of inflows,

outflows, and changes in storage over a defined time period

at a specific location. Both water cycles and budgets

consider water in solid, liquid, or gaseous form and are

typically viewed in a sequence from precipitation to

streamflow (fig. 3.1). Since most chemicals are somewhat

water soluble, the hydrologic cycle strongly influences

nutrient cycling, weathering, chemical and sediment

transport, and water quality. Furthermore, water plays vital

roles in mobilization and transport of sediment downslopes

and through stream networks.

The basic equation that describes a hydrologic budget is

Q = P - I - T - E - G - W + R + /- S ,

where

Q = streamflow,

P = precipitation,

I = interception,

T = transpiration,

E = evaporation,

G = ground water recharge,

W = water withdrawals for consumptive use,

R = return flow from outside sources, and

S = change in storage over measurement period.

Introduction

The quantity and quality of drinking water coming from a

watershed depend on processes involving surface water,

ground water, biogeochemistry, biota, atmospheric deposition, and sedimentation. Quality of drinking water supplies,

therefore, hinges on understanding of the routing of water,

its dissolved constituents, and its entrained sediments

through watersheds and ground water systems (Dunne and

Leopold 1978). Fluxes and storages of water, chemical

constituents, and sediment can be described in terms of the

average properties and variability of the cycling systems

involved. Natural forces, such as floods, can cause changes

in water flows and water quality. At times these changes can

overwhelm effects of land-use practices. It is important,

therefore, to understand natural variability of water systems

in order to have realistic expectations about the quantity and

quality of water yields from specific watersheds. Furthermore, natural variation in streamflow and water quality is

integral to the health of aquatic ecosystems (Poff and others

1997) and, thus, must be considered when attempting to

balance consumptive uses with environmental protection.

In this chapter, we provide general background information

on hydrology, dissolved constituents, atmospheric deposition, sedimentation, nitrogen impacts of surface and ground

water, cumulative watershed effects of land uses, management and policy implications, research needs, and, finally,

key points. This information is useful in assessing drinking

water issues.

1 Ecosystem Team Leader, USDA Forest Service, Pacific Northwest

Research Station, Corvallis, OR; Ecosystem Team Leader, USDA Forest

Service, International Institute of Tropical Forestry, Río Piedras, PR;

Consultant, Water Quality Management–Nonpoint Source, Eatonton, GA;

Research Plant Pathologist, USDA Forest Service, Pacific Southwest Forest

Experiment Station, Forest Fire Laboratory, Riverside, CA; Research Forest

Hydrologist, USDA Forest Service, North Central Experiment Station,

Grand Rapids, MN; and Professor of Forest Hydrology, Pennsylvania State

University, University Park, PA, respectively.
27

Units in the equation are expressed in terms of volume or

depth per-unit time (million gallons or liters per day or

inches or centimeters per year). Like all budgets, the

magnitudes of the components depend on the spatial and

temporal scale considered, and their evaluation involves

errors due to measurement and interpretation. Water budgets

for small watersheds typically have combined measurement

errors of 20 percent or greater (Winter 1981). Likewise,

municipal water that is not accounted for, that is, the

difference between the amount of water produced by

treatment plants and the amount legitimately consumed, is

commonly 20 to 40 percent of total water treatment plant

production (World Bank 1993). This difference includes

leakage from the storage and distribution system and

illegitimate uses.

For ease of communication, each component is discussed

below. However, it is important to recognize that components are tightly coupled; that is, ground water and surface

water systems, as well as atmospheric and biotic influences

on them, must be viewed and managed as a single system

and a single resource (Winter and others 1998). Many of our

past and present water-use practices and policies have

Figure 3.1—Hydrologic cycle within a watershed.

ignored these linkages. Modification in one part of the

system is likely to significantly affect other parts.

Precipitation and Atmospheric Deposition

Precipitation is often classified by physical form (liquid,

solid, or gas), size (rain, drizzle, or mist), the responsible

weather system (cyclonic, warm front, cold front, convective, or orographic, etc.), and chemistry (acidic). In general,

the amount of annual precipitation varies with elevation and

aspect relative to the prevailing winds.

Because the water molecule is dipolar and attracts other

molecules, natural precipitation contains dissolved gases in

amounts proportional to their concentrations in the atmosphere, their solubility, and ambient temperature. Uncontaminated precipitation also has low concentrations of

solutes, is slightly to moderately acidic, and has a high

redox potential. The equilibrium pH for nonsaline water in

contact with atmospheric carbon dioxide is 5.7; and rainwater and melted snow in nonurban, nonindustrial areas

typically have pH levels between 5 and 6 (Freeze and

Chapter 3

Subsurface flow system Ground water discharge

Streamflow

Exfiltration

Ground water recharge

Evapotranspiration

Overland flow

Channel flow Precipitation

Water table

Infiltration

Unsaturated

(soil moisture) zone

Saturated

(ground water) zone

Equipotential lines Flowlines
28

Watershed Processes—Fluxes of Water, Dissolved Constituents, and Sediment

Some people in the Eastern United States use cisterns,

shallow wells, or ponds for their water supply. Some sites in

coal regions can use only cistern water sources because

local ground water is extremely acidic (pH <4.0) from acid

mine drainage. The sites at risk are those where precipitation

pH is <4.5, and where surface soils do not contain enough

bases (calcium and magnesium bicarbonates) to neutralize

the precipitation acidity. Granitic bedrock, base-poor quartz

sandstones, and sandy soils derived from them have low

amounts of neutralizing bases. Basaltic rocks, sandstones

with high amounts of calcite cement, and marine, sedimentary rocks have high amounts of neutralizing bases. In areas

with acidic source water, public water supplies often adjust

pH as part of the water purification treatment, but this may

not occur in some small, private drinking water systems.

Areas in the United States where precipitation pH averages

<4.5 are restricted to the Upper Midwest and Eastern United

States (fig. 3.2). Areas with acid surface soils are in the East,

Southeast, Upper Midwest, and Northwest (fig. 3.3). These

soils correspond to areas where lakes and streams are acid,

and, thus, shallow ground water is assumed to be acid

(Church 1983).

Cherry 1979, Park 1987). In contrast, rainfall contaminated

by urban or industrial inputs can frequently have a pH as

low as 3 to 4. This acid rain is typically a result of nitrate

(NO3

-1) and sulfate (SO4-2) that are derived from the incorporation of gaseous pollutants in raindrops (Schlesinger 1997).

This increased acidity can increase the rate of weathering

and release of cations from exchange sites. Consequently,

the concentrations of metals in source water and the

corrosion of water storage and distribution systems also

increase, causing higher metal concentrations in drinking

water (McDonald 1985, Park 1987). The constituents of

concern are high acidity levels (pH < 4.5) in precipitation,

high nitrate in soil or ground water, and the interaction of

these in soil water to yield high concentrations of aluminum

(Al) and lead (Pb). In addition, highly acidic water can

dissolve lead in solder joints where copper pipes are used

for plumbing.

Atmospheric deposition is a primary source of mercury (Hg)

that can cause adverse health effects. The dangerous form,

methyl mercury, is bioconcentrated in fish that must be

eaten to endanger health. Methyl mercury in precipitation or

surface waters usually does not occur at toxic concentrations

(Nriagu and Pacyna 1988) (see chapter 2; tables 2.3, 2.4).

pH lab

≥5.3

5.1–5.3

4.9–5.1

4.7–4.9

4.5–4.7

4.3–4.5

<4.3

Sites not pictured:

Alaska 03 5.3

Puerto Rico 5.1

Figure 3.2—Average acidity (pH) of precipitation in the United States from 1988 through 1997 (National Atmospheric Deposition Program 1999).
29

Evaporation, Transpiration, and Evapotranspiration

Evaporation is the process of converting water from a liquid

or solid state to a gaseous state. Evaporation occurs from

lakes and wetlands, large rivers, soil surfaces, and accumulations of water on vegetative matter or other surfaces.

Sublimation is evaporation from snow and ice surfaces. The

rate of evaporation depends primarily on solar radiation,

temperature, wind, and the humidity gradient above the

evaporating surface. Because evaporation losses from open

water can be large, efforts have been made to reduce losses

from municipal water supplies by covering reservoirs or

storage tanks, using underground storage, controlling

aquatic growth, reducing surface area, and applying chemicals (Viessman and others 1977).

Transpiration is the process by which water is released as

vapor from plants through leaves to the atmosphere and is

influenced by soil moisture, the type of vegetation, vapor

pressure gradients across leaf surfaces, and the same factors

that affect evaporation: solar radiation, temperature,

humidity, and wind. In many cases, evaporation and

transpiration are summed and reported as one process,

termed evapotranspiration (ET). Unlike other pathways, ET

returns water to the atmosphere without solutes and, thus,

increases the concentrations of solutes in the water remaining in terrestrial or aquatic systems.

Interception, Throughfall, and Stemflow

Interception is the process whereby precipitation collects on

vegetation and evaporates instead of falling directly or

indirectly to the ground. Throughfall is water that may or

may not contact vegetation as it passes through the vegetative canopy and eventually falls to the ground. Stemflow is

water that reaches the ground flowing along the stems of

vegetation. The amount of interception varies with the

magnitude and intensity of rainfall, the structure and

composition of the canopy, the season, and the form of

precipitation (Anderson and others 1976). In general, forest

vegetation intercepts more precipitation than grasslands, and

conifers intercept more water than hardwoods.

Chapter 3

Figure 3.3—The acid buffering capacity (total alkalinity) in lake water and, by inference, the surrounding water in soils and ground water for the

United States. Where lakes have low or negative alkalinities in the black areas, surface water pH values may be <4.5. (Map prepared by J.M.

Omernick, G.E. Griffith, J.T. Irish, and C.B. Johnson with the U.S. Environmental Protection Agency.)

Total alkalinity

(μeq L )

<100

100–400

>400

-1
30

Watershed Processes—Fluxes of Water, Dissolved Constituents, and Sediment

Solutes in throughfall and stemflow consist of both new and

recycled constituents. New inputs are chemicals and

particulates that originated outside the area of interest and

collect on vegetative surfaces through a process called dry

deposition. Recycled constituents come from decomposition

and leaching of plant tissue. In general, the cation and anion

concentrations in throughfall are 2 to 100 times those of

rainfall (Wenger 1994). In deciduous forests, throughfall and

stemflow have the highest concentrations during the

summer when the forest has the largest leaf area.

Soil Water

Once water passes through the vegetative canopy, it comes

in contact with the forest litter layer and soil surface, where

it either infiltrates, evaporates, temporarily ponds, or leaves

the area as surface runoff. Water that infiltrates can reside in

many subsurface areas (fig. 3.4) and remain below the

surface for a period ranging from seconds to millennia. The

rate at which water enters the soil—called the infiltration or

percolation rate—is influenced by the magnitude and

intensity of rainfall, the type and extent of vegetation cover,

and the temperature and condition of the surface. In general,

the amount of infiltration in a watershed decreases with the

amount of pavement and increases with forest cover and soil

organic matter content.

Ground Water

As water flows through the upper soils and ground water

system, it interacts with its surroundings and undergoes

chemical changes. Typically, organic acids are produced and

nutrients are immobilized in the upper layers of fresh forest

litter (Schlesinger 1997). Mineralization of nitrogen (N),

phosphorus (P), and sulfur (S) is usually greatest in the

lower forest floor and upper mineral soil. As water passes

through these layers, organic acids and other decomposition

products can produce undesirable odors or taste and can

increase water hardness (Freeze and Cherry 1979).

Figure 3.4—Schematic illustrating ground water terms and concepts.

Water

table

Potentimetric

surface

Flowing

artesian

well

Gaining

stream

Seep

Aquitard

Ground water

(phreatic water)

Water

 table

Bedrock

Land

surface

Water table

well

Capillary

fringe

Confining bed

Unconfined aquifer

Spring

Confining

bed

Losing

stream

Perched water

table and aquifer

Confined

aquifer

Vadose zone

Zone of

saturation
31

Exchanges between ground water and mineral particles

generally increase the concentrations of total dissolved

solids (TDS), cations, calcium bicarbonate (Ca(HCO3)

2

),

magnesium bicarbonate (Mg(HCO3)

2

), calcium sulfate

(CaSO4

), and magnesium sulfate (MgSO4). Microbial

processes are usually responsible for methanogenesis,

denitrification, sulfate (SO4

-2) and ferrous iron (Fe+3)

reduction and the breakdown of natural and synthetic

organic compounds. Denitrifying bacteria have been

collected from depths of 1,000 feet [300 meters (m)] and

sulfate-reducing bacteria can remove most sulfate within 50

feet (15 m) of the soil surface (Schlesinger 1997). The travel

time and distance needed to remove viruses or synthetic

organics vary considerably, but typically are on the order of

days to years and feet to miles (meters to kilometers).

Because of these exchanges and microbial processes, older

and deeper ground water generally has greater hardness but

fewer organic pollutants and often needs less treatment than

surface water or shallow, young ground water.

Streamflow

Streamflow is broadly divided into two types, stormflow

and baseflow. Water flows to streams by three processes:

(1) overland flow (or surface runoff), (2) interflow (or

subsurface stormflow), and (3) ground water flow (Linsley

and others 1982). Overland flow involves water that travels

over the ground surface to a stream channel. Interflow

involves water that infiltrates into the upper soil layers and

moves laterally until it enters a stream channel. In most

forested watersheds, the rate at which water can infiltrate

into the soil is greater than the rate of rainfall. Therefore,

overland flow is relatively rare or is limited to areas with

shallow, degraded soils or saturated areas in a watershed. In

contrast, interflow is common, especially in areas with thin,

porous soils that become saturated during storms or in areas

where subsurface soil pipes or macropores have developed.

These subsurface conduits can have diameters that range

from fractions of an inch (centimeter) to several feet

(meters). In some instances, the flow velocities within pipes

are sufficient to cause in-situ erosion. Subsurface conduits

can eventually become so large that they collapse and form

incipient stream channels. Because water in these pipes is

rapidly transferred to streams, the purification that commonly occurs as water slowly travels through microscopic

soil pores does not take place.

The chemical characteristics of streamwater depend on its

source and the flow path and transit time to the stream. In

general, the concentrations of dissolved solids decrease with

increasing discharge and increase with the length of the flow

path and the amount of time the water has traveled across

the landscape. These generalizations are especially true of

highly soluble and typically nonbiologically limiting ions—

like calcium (Ca+2), magnesium (Mg+2), sodium (Na+1),

silica (SiO2), chloride (Cl-1), bicarbonate (HCO3

-1), and

sulfate SO4

-2 —associated with chemical weathering

(Schlesinger 1997). They are also generally true for chemicals derived from point sources that enter streams at

relatively constant rates. In contrast, the concentrations of

sediment and particulate matter derived from physical

detachment and chemicals derived from the flushing of the

land surface or shallow subsurface tend to increase with

stream discharge and the proportion of surface or storm

runoff in the stream.

Once in the stream, constituents may be transported in

solution, in suspension, or attached to particles. Metabolic

activity in a stream depends on upstream inputs, internal

(algae, aquatic plants), and external (leaves, dissolved

organic carbon) sources of food and nutrients. The major

processes affecting dissolved oxygen (DO) in a stream are

reaeration, carbonaceous and nitrogenous deoxygenation,

sediment oxygen demand, and plant photosynthesis and

respiration (Marzolf and others 1994, Newbold and others

1982, Vannote and others 1980). In streams with large,

standing crops of submerged aquatic plants, the uptake of

carbon dioxide during photosynthesis can remove enough

carbonic acid from water to increase daytime pH by several

units.

From headwaters to lowlands, streams change in their

morphology, water chemistry, and biotic communities

(Vannote and others 1980). In general, headwater streams

are shaded by terrestrial vegetation and have biotic communities that depend on leaf litter and other natural terrestrial

sources of organic matter. The water in these streams also

tends to have low concentrations of TDS. In downstream

areas, the amount of light entering the channel, the contribution of ground water, and anthropogenic contaminants

generally increase. Consequently, TDS and contaminants

tend to increase and aquatic plants and algae rather than

terrestrial plants become the major source of organic matter

inputs to streams. As dissolved constituents are transported

downstream, they are converted to organic forms and

accumulated in organisms until they die and are recycled.

This change between organic and inorganic forms may

occur several times as nutrients “spiral” down the channel

from the headwaters to the lowlands (Newbold and others

1982).

The physical and biotic changes in water quality that occur

along a river can also affect the operation and cost of

municipal water treatment. Moreover, because concentrations of TDS and pollutants increase downstream, water

withdrawn from lower reaches of streams typically needs

Chapter 3
32

Watershed Processes—Fluxes of Water, Dissolved Constituents, and Sediment

more treatment than water from undisturbed, forested

headwater areas. Nevertheless, because they usually have

greater volumes of water and less seasonal variability in

supply, lowland water intakes are often more reliable

sources of water.

Water Withdrawals and Return Flow

Water withdrawal is the process of removing water from a

hydrologic system and conveying it to a place for offstream

use. Nonwithdrawal or instream uses include navigation,

hydropower generation, recreation, and the maintenance of

aquatic habitat. Return flow includes water that is added to a

hydrologic system after it has been withdrawn and used or

leakage from storage and distribution systems. In municipal

watersheds, return flows are typically from point discharges

from sewage treatment plants, irrigation systems, or

industrial sources. However, in some urban areas, nonpoint

discharges and conveyance losses from leaky pipes or

irrigation ditches may contribute significant volumes of

return flow.

The influence of return flow on water quality is a function

of the quality and quantity of the return flow, the quality of

the receiving waters, and the distance below the discharge

point and turbulence of flow in the receiving waters. In

general, a streamflow to wasteflow ratio of about 40 to 1 is

needed to safely dilute most raw, untreated waste (Gupta

1995). A ratio of about 2 to 1 is needed to dilute waste from

most secondary water treatment plants. Unless the effluent is

disinfected, most wastewater treatment does not markedly

reduce pathogens. In the United States, most effluent is

disinfected with chlorine. Excessive chlorination, however,

may lead to toxicity problems for aquatic organisms in

receiving waters. See chapter 5 for more discussion of

withdrawals and return flows on drinking water quality.

Effects of Nitrogen Deposition on Stream

and Ground Water Quality

An increasing number of studies from wildland watersheds

in many parts of the World demonstrate a link between

chronic nitrogen (N) inputs from air pollution and nitrate

levels in streamwater and ground water emanating from

these watersheds (Fenn and Poth 1999, Stoddard 1994).

Nitrogen saturation is the term now commonly applied to

the phenomenon of ecosystems, which export high nitrogen

levels as a result of available nitrogen in excess of biotic

demand and of watershed nitrogen retention capacity (Aber

and others 1989). The excess nitrogen is predominantly

exported as nitrate in drainage waters, but gaseous losses of

nitrogen from soil and in the riparian zone can also be

important. The source of the excess nitrogen is usually

elevated nitrogen inputs from the atmosphere (nitrogen

deposition), but nitrogen fertilizer application and nitrogenfixing plant species, which convert free dinitrogen gas (N2

)

in the atmosphere into organic forms of nitrogen in plant

tissue, are other sources of excess nitrogen.

Surface water and ground water are commonly contaminated with elevated nitrate in nitrogen saturated watersheds

(Berg and Verhoef 1998, Fenn and Poth 1999). However,

watershed-level studies of nitrogen saturation tend to focus

on nitrate concentrations in streams, which is generally

easier to access than ground water. Ground water can be

sampled from wells or from ground water-fed springs where

they occur. In many instances, streamflow originates from

springs. If the watershed is nitrogen saturated, stream nitrate

may come from contaminated ground water. For example,

hydrologic studies in the nitrogen-exporting Neversink

River watershed in the Catskill Mountains, NY, found that

during the summer low-flow period, streamflow originated

from perennial springs. The springs discharged deep ground

water that was recharged during the dormant season 6 to 22

months earlier when soil nitrate levels are highest (Burns

and others 1998). During the summer, nitrate concentrations

in these streams were higher than in shallow ground water,

which was recharged during that growing season when

plants take up nitrogen and because the streamwater

originated from deep ground water. In the summer-dry

climate of southern California, nitrate concentration in

springs was an excellent indicator of watershed nitrogen

status (Fenn and Poth 1999). Nitrate concentrations in

springwater did not vary seasonally, suggesting that

springwater or ground water may be a more useful indicator

of nitrogen saturation due to its greater temporal stability

compared to surface runoff.

Water from forested watersheds is commonly used to

improve drinking water quality by blending it with lower

quality water from other sources. Water from nitrogensaturated watersheds has high nitrate concentrations that

only exceed the Federal drinking water standard following a

major disturbance, such as fire, harvesting, etc. (Riggan and

others 1994). Water from nitrogen-saturated watersheds has

reduced dilution power, leaving water resource managers

with the need to implement much more expensive water

treatment options. A high nitrate level in drinking water is

an important human health concern (see chapter 2). In

addition, excess nitrate exported to lakes and estuaries

contributes to eutrophication of these bodies.

The geographic extent of nitrogen-saturated watersheds in

North America (Fenn and others 1998, Stoddard 1994) is

only partially known, largely because research on this topic
33

in North America is still in the relatively early stages.

Nitrogen saturation cannot be predicted based solely on the

amount of nitrogen deposition. The rate of total nitrogen

deposition is a factor contributing to nitrogen saturation.

However, for most areas, only nitrogen deposition in rain

and snow have been measured. Dry deposition of nitrogen

in gaseous or particulate forms has been measured in

relatively few places because measuring dry deposition is

still in the experimental stage of development. In areas of

high air pollution, especially in dry climates, dry deposition

may be a large contributor of nitrogen to forests.

The percentage of forest land cover in North America

exhibiting severe symptoms of nitrogen saturation, such as

large nitrate export losses, is relatively low. Much larger

areas of forested lands exhibit moderate nitrate export, a

sign that they may be vulnerable to nitrogen saturation in the

future. Forest production may be enhanced in some of these

areas as a result of the inadvertent atmospheric nitrogen

fertilization (Fenn and others 1998). However, the problem

of excess nitrogen is not trivial. Signs of nitrogen saturation

have been reported in the headwater streams in forests in the

Catskill Mountains (Murdoch and Stoddard 1992) and the

Transverse Ranges in southern California, both of which

supply drinking water to millions of inhabitants (Fenn and

Poth 1999, Riggan and others 1994).

In many nitrogen-saturated watersheds, most of the nitrate

leached from the ecosystem is cycled through plant litter,

organic matter, and microbes prior to being exported.

Although some studies suggest that nitrogen deposition

above a threshold level can eventually lead to elevated

nitrate loss in temperate forests (Dise and Wright 1995),

there are also clear exceptions to the pattern. The relationship between nitrogen deposition and nitrate leaching is

confounded by complex, nitrogen-cycling processes and the

biological and physical characteristics of forested watersheds. Ecosystem controls on nitrogen processing and

nitrogen loss are poorly understood at the mechanistic level.

Thus, our ability to predict nitrogen losses from watersheds

exposed to chronic atmospheric nitrogen deposition is

limited. Plant and soil indicators of ecosystem nitrogen

status are available, and they can be used to monitor and

identify ecosystems for symptoms of nitrogen saturation

(Fenn and Poth 1998). Suggested indicators include litter

carbon-to-nitrogen ratio, foliar nitrogen-to-phosphorus or

nitrogen-to-magnesium ratio, and ratios of rates of soil

nitrification-to-mineralization.

Although our knowledge is incomplete of how different

ecosystem types process nitrogen, certain characteristics that

are known to predispose ecosystems to nitrate loss can be

used to identify watersheds at risk of elevated streamwater

nitrate concentrations (Fenn and others 1998). Such factors

include steep slopes and coarse-textured, shallow soils, or

both that encourage rapid runoff with little opportunity for

biological uptake and retention of dissolved nitrate. Mature

forest stands may have large stores of organic nitrogen in

the soil, the forest floor litter layers, and the old trees. Plant

nitrogen demand from the soil is lower in old stands than

vigorously growing younger stands with lower ecosystem

nitrogen stores. Older stands, therefore, are particularly

prone to nitrogen saturation, even where rates of atmospheric nitrogen deposition are low to moderate (Foster and

others 1989).

Forest type may influence ecosystem susceptibility to

nitrogen saturation. Recent studies suggest conifer stands

are more prone to nitrogen saturation and nitrogen loss than

hardwood stands (Aber and others 1995). Preliminary

results indicate elevated nitrogen inputs may convert some

conifer stands to deciduous forests with high nitrogen

cycling rates (McNulty and others 1996).

Low soil cation capacity may predispose forests to symptoms of nitrogen saturation if other macronutrients, such as

calcium or magnesium, become limiting. This can result in

nutrient imbalance in some plants, disruption of plant

function, forest decline, decreased nitrogen demand, and

increased nitrate leaching (Durka and others 1994).

High elevation ecosystems, which include some class I

wilderness areas in national forests and national parks, are

especially prone to high nitrate losses, even where atmospheric nitrogen deposition is moderate. High elevation

systems are often characterized by steep slopes, coarsetextured soils, exposed bedrock, and sparse vegetation with

low plant nitrogen demand. Low temperatures also result in

reduced plant and microbial nitrogen retention. Nitrate

runoff in these systems is particularly high during high

runoff periods, such as during spring snowmelt and after

large storms. High elevation tundra ecosystems in the Front

Range of the Colorado Rockies are nitrogen saturated with

low-to-moderate nitrogen deposition rates (Williams and

others 1996).

Since the norm for most watersheds of the Northern

Hemisphere is nitrogen limitation rather than nitrogen

excess, land managers have little experience dealing with

the problem of nitrogen saturation. The causes and effects of

nitrogen saturation are areas of active research. The generalizations discussed in this section are supported by many

recent studies, but little if any research in North America has

focused on the effectiveness of silvicultural treatments for

reducing high nitrate concentrations in runoff water.

Previous studies on the effects of fire, harvesting activities,

Chapter 3
34

and other silvicultural treatments on nitrate runoff provide

clues as to possible management options in nitrogensaturated watersheds. Further research is needed, but it

seems likely that increasing plant nitrogen and water

demand by encouraging the growth of young, fast-growing

deciduous forests, which are increasing in biomass, is likely

to reduce nitrate runoff. Other promising strategies for

reducing nitrate in runoff focus on the riparian zone.

Reducing the amount of nitrogen stored in the ecosystem is

another strategy for reducing high nitrogen losses in runoff.

For example, forest harvest intensity affects the amount of

nitrogen left in the system and, thus, the amount of nitrogen

that can leach from the watershed. Whole-tree harvesting

was found to reduce the amount of nitrate in runoff compared to less intensive harvests in which slash was left in the

forest after the harvest (Hendrickson and others 1989).

Removing slash in harvest operations not only reduces the

amount of nitrogen in the system, it also allows for more

rapid regeneration of vegetation following the harvest,

resulting in greater vegetative nitrogen demand and nitrogen

retention. However, in Eastern North America, cation

depletion in soil is thought to be a serious threat to forest

sustainability and productivity in some areas (Federer and

others 1989). The more intensive harvesting regimes would

likely exacerbate this problem, and could only be used

effectively if fertilizer is applied to replace the limiting

nutrients. In fact, if a forest is growth-limited by a nutrient

other than nitrogen, e.g., phosphorus, sulfur, or calcium,

fertilizing with that nutrient will likely increase plant

nitrogen demand and should result in lower levels of nitrate

runoff (Stevens and others 1993). Nitrogen fertilization

should be avoided in forests showing signs of nitrogen

saturation because it is likely to exacerbate nitrate levels in

runoff. There is evidence that some plant species are

associated with more rapid rates of nitrate production

(nitrification) and, thus, increase the risk of elevated nitrate

runoff. Replanting with species with lower nitrification rates

and greater nitrogen consumption or both and storage rates

is another option for reducing nitrogen-saturation effects.

Use of this approach, however, will require information on

the nitrogen-cycling properties of the tree species under

consideration.

Prescribed burning may serve a similar function to harvesting in removing organic nitrogen stores and stimulating

more vigorous vegetation growth after burning. It has been

proposed as a management alternative in nitrogen-saturated

watersheds (Riggan and others 1994). However, nitrate

concentrations may increase dramatically for a time after

burning in nitrogen-saturated sites, and care must be used to

avoid erosion and high sediment transport. Over the long

term, however, nitrate concentrations are expected to

decrease following moderate burns. More research will be

needed to determine if this approach is effective in different

ecosystem types. Even if prescribed fire is shown to be

effective in reducing nitrate runoff, other political, sociological, logistical, environmental, and economic restraints

can sometimes make this approach difficult to implement.

In many cases, the best opportunity for reducing nitrate

concentrations in runoff will likely be in the riparian zone,

where nitrogen cycling is particularly dynamic. Nitrate

levels can be reduced as nitrogen is taken up by riparian

vegetation or by aquatic biota. The other major mechanism

for reducing nitrate levels is denitrification, which is the

conversion of nitrate to gaseous forms of nitrogen by a

specialized group of anaerobic microorganisms. Buffer

strips of riparian vegetation can be managed for maximum

nitrogen retention and as a carbon source for denitrifying

bacteria. Buffer strips 15 to 100 feet (5 to 30 m) wide have

been shown to be highly effective in nutrient retention in

surface runoff and in subsurface flow (Haycock and others

1993). Wetlands can also serve as effective nitrogen sinks,

and restoration or creation of wetlands is another option for

management of high nitrate runoff to coastal areas

(Fleischer and others 1991). Although some active management practices have the potential to reduce impacts of

nitrogen-saturation on drinking water, none have been tested

in nitrogen-saturated watersheds. Active management

options may be limited or inappropriate in areas such as

alpine zones or wilderness.

Sediment Production and Transport

Sediment is moved from slopes to stream channels and

through stream networks by a great variety of processes.

Some of these processes are pervasive and persistent, such

as the removal of fine-grained weathering products in

suspension. Other processes operate infrequently and even

catastrophically, as in the case of rapid landslides. Sediment

transport through stream systems involves a variety of

processes ranging from transfer of dissolved material, to

movement of fine particulate material in the water column,

to rolling of coarse particles along the streambed. Thus, the

movement of these materials through a watershed involves a

series of linked transfer processes and storage sites, such as

gravel bars and floodplains. As with hydrological and

biogeochemical cycling, the routing of sediment through

watersheds has both long-term, average properties and very

significant fluxes during extreme events.

Small sediment [<0.06 millimeters (mm), silt size] tends to

move relatively rapidly through the channel system as wash

load. Fine sediment is a major cause of turbidity. Larger

Watershed Processes—Fluxes of Water, Dissolved Constituents, and Sediment
35

sediment moves as bed material load and can have long

residence times. Bunté and MacDonald (1999) comprehensively reviewed the literature dealing with sediment transport distance as a function of particle size. Travel distance

for suspended load (wash load plus some sand) ranges from

1.2 to 12 miles [2 to 20 kilometers (km)] per year, whereas

bed load (pebbles and cobbles) travels only 0.012 to 0.3

miles (0.02 to 0.5 km) per year. In low-gradient channels,

such as those found in portions of the Lake States and the

Southeastern United States, residence times for sands can

range from 50 to 100 years (Phillips 1993, Trimble 1999).

Studies in the Western United States show sediment storage

times in active stream channels ranging from 5 years to

hundreds of years, depending on particle size and the type of

sediment deposit (Madej and Ozaki 1996, Megahan and

others 1980, Ziemer and others 1991).

Effects of floods, landslides, and chronic processes on

sediment production have been widely studied and are

highly relevant to evaluating the effects of forest and

rangeland management on drinking water supplies. However, little work has linked results of sedimentation studies

directly with the quality of drinking water. The relevant

approaches to studies have included small watershed

experiments (Binkley and Brown 1993, Fredriksen and

others 1975, Likens and Bormann 1995, Swank and

Crossley 1988), landslide inventories (Sidle and others

1985), sediment budget analyses (Reid and Dunne 1996,

Swanson and others 1982), magnitude-frequency analysis

(Wolman and Miller 1960), and studies directly targeting

water-quality issues for particular storm events (Bates and

others 1998). The latter type of study is most germane to our

topic here but commonly resides in the gray literature and

consulting reports. The other study approaches listed

commonly present results in terms of annual or longer time

scales because they typically address questions of soil loss,

nutrient balances, and landscape denudation, rather than

drinking water quality where problems typically develop on

the time scale of individual storm events.

The capacity of watersheds with near-natural vegetation to

produce sediment that reduces drinking water quality

depends on soil properties, topography, climate, and

vegetation conditions. Steeper slopes, of course, favor

sediment production. Certain rock and soil types are prone

to landslides (Sidle and others 1985) and to produce

distinctive clay minerals that can cause persistent turbidity

(Bates and others 1998, U.S. General Accounting Office

1998, Youngberg and others 1975). Effects of climate are

complex. More precipitation favors water-driven erosion

processes, but wetter conditions also favor vegetation

development. Vegetation suppresses soil erosion by developing a litter layer that protects the soil from surface erosion

and by developing root systems that contribute to soil

strength.

Numerous studies in steep, unstable mountain land have

documented that a substantial share of long-term sediment

production occurs during extreme events, particularly when

landslides are triggered (Swanson and others 1987).

Inventories of small, rapid landslides reveal that these

natural processes occur in forested terrain, as well as in

areas disturbed by land management activities (Sidle and

others 1985). Large, slow-moving landslides, commonly

termed earthflows, are also natural, and, in some cases, they

persist for millennia. Earthflow areas may be more prone to

produce persistent turbidity because the montmorillonite

clays that degrade water quality also cause the slow,

creeping deformation characteristic of this type of landslide

(Taskey and others 1978). Earthflows slowly encroach on

stream channels, constricting them over periods of years.

Then, floodwater undercuts the toe of the earthflow, causing

streamside slides and delivering turbidity-producing

sediment.

Interactions among geomorphic processes can increase the

availability of sediment for many years. Major floods can

deliver massive quantities of sediment to channels often by

initiating landslides. For many years afterwards, suspended

sediment loads may be unusually high during storms. In

these cases, large amounts of sediment build up in transient

storage sites along the stream, where they are mobilized by

subsequent storms (Brown and Ritter 1971). Large sediment

input to rivers causes channel aggradation, widening, and

lateral cutting into floodplain deposits and toes of hillslopes,

thus entraining stored sediment. In some cases, stored

sediment and colluvium may have weathered sufficiently to

contain clay minerals that cause high levels of turbidity.

Thus, a major flood can affect erosion and sediment

transport processes during interflood periods. These

processes are more evident in areas of extreme sedimentation (Kelsey 1980), but these interactions probably operate

less conspicuously in systems with lower overall rates of

sediment input.

North Santiam River Case Study

Many of these complex interactions among natural processes, land use, water management, and drinking water are

exemplified by the case of the city of Salem, the capital of

Oregon. High levels of turbidity led Salem to temporarily

suspend use of its drinking water treatment facilities that

draw water from the 766-square mile (1960-km2

) North

Santiam River Basin during a major flood in February 1996

(Bates and others 1998). The x-ray diffraction analysis of

suspended sediment in the turbid water revealed smectite

clay, which forms exceedingly small particles with surface

Chapter 3
36

electrical properties that permit them to remain in

suspension for many weeks. Using the clay mineral analysis,

it is possible to “fingerprint” large, slow-moving landslides

locally termed earthflows as a major source of turbidityproducing smectite (Bates and others 1998). Thus, natural

geomorphic features (earthflows) and processes (earthflow

movement and flooding, eroding earthflows) play a strong

role in the elevated turbidity. The degree to which current

land-use practices affect earthflow movement and the floods

eroding the toes of these ancient landslides are debatable.

These relations between rock and soil types or both,

processes of sediment delivery, and downstream water

quality are common in other areas of the Cascade Range in

Oregon (Taskey and others 1978, Youngberg and others

1975), and the general approach to fingerprinting causes of

water-quality degradation can be applied more broadly.

In addition, a large flood-control reservoir in the middle of

the North Santiam watershed, which, while reducing flood

levels downstream, releases turbid water over a period of

many days, thus exacerbating water-quality problems. As

the city of Salem moved to increase chemical treatment of

water from the North Santiam River, computer chip manufacturers expressed concern that the introduced chemicals

would degrade water quality from the perspective of their

uses.

Geographic and temporal variation in watershed response to

floods and land use is great, as are the implications for

drinking water supplies (U.S. General Accounting Office

1998). While Salem’s water treatment system was temporarily shut down due to high turbidity levels, more advanced

treatment facilities, such as those of Eugene, OR, were

treating water with higher turbidity (U.S. General Accounting Office 1998). However, in the generally stable watershed supplying Portland, OR, a wet winter triggered a

single, natural landslide in an unmanaged area that interrupted water supplies because of high turbidity levels.

Logging and roads in the watershed have been controversial,

but it has been difficult to demonstrate they have degraded

water quality. Watersheds with extensive areas of unstable

rock and soil types are likely to have lower water quality,

even if land-use activities were absent.

Natural Disturbance Processes

Natural processes that severely disturb vegetation, such as

fire and extensive wind toppling of forests, can affect

drinking water quality. Windstorms in the Eastern United

States range in scale from localized storms (Hack and

Goodlett 1960) to regional hurricanes (Foster and others

1997). The potential of wildfire to degrade drinking water

supplies is a prevalent problem in western mountain

landscapes, where fire strongly affects both pulses and longterm patterns of sediment production (Swanson 1981) as

well as nitrogen concentrations in runoff (Beschta 1990).

Fire is also prevalent in grassland systems, but its effects on

sediment production can be quite limited if the fire does not

kill the vegetation or change the roughness of the ground

surface (Gray and others 1998: 162) (see chapter 12).

Effects of these vegetation disturbances on downstream

water quality depend on the severity of disturbance to

vegetation and soil, the timing of precipitation in relation to

vegetation disturbance, and the propensity of the landscape

and ecosystem to produce compounds that degrade water

quality. However, we know of no studies directly addressing

drinking water quality in response to these processes.

In many regions of the country, streams are currently

transporting sediment from past land uses and management

practices as well as sediment from past catastrophic events,

such as wildfires, large storms, and landslides. The rate of

transport depends on the size of sediment particles, gradient

of streams, and streamflow. At many locations, sediment

from past erosion is influencing present-day channel

conditions and sediment transport. In several regions, forests

were cleared for grazing, mining, and agriculture in the

1800’s and early 1900’s. For example, forests in the Piedmont region of the Southeast were cleared for agriculture

and were abusively treated, causing large increases in

erosion (Trimble 1969, 1974). The excessive sediment

supply exceeded the transport ability of the streams. Huge

volumes of sediment were deposited in the stream channels

and floodplains. The severely eroded fields were eventually

abandoned and reverted naturally to forest or were planted

to trees and pasture under conservation programs in the midto-late 1900’s. The landscape stabilized and sediment yields

to streams were greatly reduced. Because the runoff from

the landscape carried little sediment, the streams had more

energy available to transport sediment and began transporting sediment released from floodplain storage as the streams

have cut downward and headward through the stored

sediment (Trimble 1999). The process continues today in

many river systems. For these streams, much of the sediment being transported today is from long-abandoned land

uses.

Several issues and risks that may result from sediment

transport from past and abandoned land uses include

(1) sediment yields from a watershed may be higher than

expected from present forest and grassland management.;

(2) streams remobilizing stored sediment often have

unstable channels and banks; and (3) stored sediment from

past land uses may contain chemicals and metals that impair

water quality.

Watershed Processes—Fluxes of Water, Dissolved Constituents, and Sediment
37

Cumulative Watershed Effects

The National Environmental Policy Act (NEPA) of 1969

stipulates that cumulative effects must be considered in

evaluating environmental impacts of proposed Federal

projects. To implement this legislation, the Council on

Environmental Quality (CEQ Guidelines, 40 CFR 1508.7,

issued 23 April 1971) provided the relevant definition:

Cumulative impact is the impact on the environment

which results from the incremental impact of the action

when added to other past, present, and reasonably

foreseeable future actions . . . Cumulative impacts can

result from individually minor but collectively significant actions taking place over a period of time.

A cumulative watershed impact influences or is influenced

by the flow of water through a watershed (Reid 1998).

Cumulative watershed effects, a phrase which has widely

replaced reference to “impacts,” can be additive or synergistic and involve modification of water, sediment, nutrients,

pollutants, and other watershed system components. An

example of such effects would be where forest roads and

timber cutting contributes to increased peak streamflows

and sediment loads, leading to aggradation of downstream

areas, which in turn results in lateral channel migration

causing streambank and floodplain erosion, which entrains

additional sediment.

Reid (1993) provides a broad and detailed summary of

cumulative watershed effects of diverse land-use activities,

such as grazing, roads, logging, recreation, and water

extraction. She also addresses alternative approaches for

assessing cumulative effects (Reid 1993, 1998). Cumulative

effects can be addressed by examining the changes triggered

by a particular land-use activity and how these changes

interact with effects of other land uses and natural processes.

Such an approach is best undertaken as a long-term study

with substantial focus on mechanisms of transport, transformation, and storage within the watershed. An alternative

approach is to work backward from a detected impact and

attempt to interpret the chain of events and processes

responsible. Each approach has strengths and weaknesses.

An important development in anticipating and hopefully

minimizing cumulative watershed effects has been the

watershed analysis developed for use by Federal (e.g.,

Regional Ecosystem Office 1995) and State (Washington

Forest Practices Board 1995) agencies in the Pacific

Northwest. The general objective of the Federal watershed

analysis procedure is to gain an understanding of present

and prospective future mechanisms affecting watershed

conditions. Thus, watershed analysis provides a useful

starting point for assessments of cumulative watershed

effects. However, Reid (1998) asserts that neither of these

“widely used watershed analysis methods provides an

adequate assessment of likely cumulative effects of planned

projects.”

See appendix C for a case study on the cumulative impacts

of land use on water quality in a Southern Appalachian

watershed. Watershed analysis is in an early stage of

development and application. It recognizes that water supply

and watershed management issues must be addressed from

an interdisciplinary, whole-system perspective. Although

watershed analysis may provide a useful first step for

assessing how multiple, simultaneous forms of management

affect sources of drinking water, there is a need to develop

better models to predict watershed cumulative effects.

Management and Policy Considerations

Existing information on the hydrologic cycle and sediment

routing systems is generally good in terms of understanding

natural controls on water flow and quality. This knowledge

is based in part on a long history of water use, detection of

problems, and studies to build a basis for problem solution

(Anderson and others 1976, Binkley and Brown 1993).

Long-term studies in experimental watersheds, including

control watersheds, give a lengthening record of variability

in water quality; but records seldom include the instantaneous sample concentrations that are most useful in addressing questions about drinking water quality. These and other

long-term and short-term studies generally corroborate

results of earlier work.

Despite our growing knowledge of natural patterns of water

flows and quality, new land management practices are

stretching the reliability of existing information. For

example, long-term studies of water quality from small

watersheds involve forest land-use treatments that are unlike

those being used today. These new practices involve lower

intensities of site treatments, e.g., partial cutting vs.

clearcutting, lower intensities of slash fires, and longer

rotations, so the treated and control watersheds in experimental watershed studies bracket the conditions created by

newer treatments, providing a basis for estimating effects.

Also, some new management treatments are aimed at

ecosystem and watershed restoration, which may include

use of fire in fire-prone systems where fire has been

excluded for many years. Reintroduction of fire into forests

where it has been suppressed for many decades will require

evaluating short-term risks of degraded water quality against

the expectation of reducing longer term risks of highseverity wildfire resulting from higher fuel loads (see

chapter 12). In these cases, water-quality objectives will

compete with other ecological and management objectives.

Chapter 3
38

Furthermore, water-quality standards are changing for a host

of reasons, not only for drinking water use, but also to meet

refined ecological objectives such as protection of threatened and endangered species and to supply high-technology

companies, which may not want water subjected to the

standard chemical treatments for drinking water. These

factors, in the evolving social and biophysical environment

of drinking water issues, indicate the importance of explicitly revealing the limits of knowledge and possibly taking a

risk assessment perspective in addressing emerging drinking

water issues.

Because present knowledge pertains to the specific geophysical and biological conditions of study sites, we have

limited ability to extrapolate findings more broadly. However, various efforts to develop regional and national spatial

data bases on soil, vegetation, and topography in relation to

watersheds supplying drinking water are building a basis for

extrapolating findings across much larger areas (Hunsaker

and others 1992). These data compilation efforts are a

common factor in many bioregional assessments (Johnson

and others 1999).

Important challenges are emerging in cases where competing objectives call for integrated understanding of ecological, geophysical, and human factors over large watersheds.

Bases for carrying out this integration are being developed

in watershed analyses conducted in a variety of contexts,

including dam relicensing procedures under Federal Energy

Regulatory Commission and in the Northwest Forest Plan in

the Pacific Coast. These large-scale, integrative assessments,

which form the basis for addressing management and policy

issues around major water supplies, are substantially

advancing knowledge.

Research Needs

1. Studies are needed of key hydrological, biogeochemical,

and sediment transport processes that affect drinking

water quality. Research needs include (a) development of

reliable methods to analyze routing of these materials

through watersheds; (b) determining the chemical

processes associated with sediment in transport and

storage; and (c) refining understanding of the roles of

past and present land-use practices on water quality and

sediment production, including land-use-related sediment

released from long-term storage. The target processes

may vary among ecological, geological, and climatic

settings across the country.

2. Better understanding is needed of the overall cycling and

routing of water, dissolved constituents, soil, and

sediment in natural and managed watersheds. Studies to

gain this understanding need to be framed so that

questions such as the following can be addressed: How

has management of ecosystems and water systems altered

natural, historical water flow regimes, biogeochemistry,

and sediment routing? How have the types and degrees of

these past and prospective future alterations of these

systems altered their ability to meet objectives for water

supplies, ecosystem health, and other goods and services?

How might climate change alter these systems?

3. Watershed-scale assessments are needed of water

pollution and sediment sources operating during and after

extreme events. It is important to better quantify the

significance of these events by maintaining long-term

studies, by monitoring the quality of source water at

drinking water treatment facilities, and short-term,

intensive studies targeting effects of particular storm

events. While many of these assessments are conducted

by management agency personnel, there is a continuing

need for participation by researchers to foster development of science at this geographic scale and scope of

interdisciplinary work.

4. There is need for integration of information from specific

watershed studies to broad-scale management applications. This sort of regional analysis is occurring in a

variety of management and research sectors on topics

relevant to drinking water quality. Relevant tools, such as

Geographic Information Systems, analytical approaches,

and data bases are available.

5. Good records of raw and treated water at treatment

facilities would provide researchers and others with much

improved data bases for evaluating long-term trends in

water quality from watersheds used as drinking water

sources. Existing records should be examined for trends

in water quality. Though not a research need itself, this is

an important step in ultimately furthering research into

causes and cures of water-quality problems.

6. Management options for controlling streamwater nitrate

levels need to be tested for efficacy. Examples of options

include tree harvesting; planting more rapidly growing

and nitrogen demanding species; thinning, or other

vegetation management approaches; prescribed burning;

fertilizer application; and vegetation buffer strips in the

riparian zone. More research is needed on vegetation type

or species’ effects on nitrification, since nitrate production rates are key in controlling nitrate losses. Information is also needed on tree species with the capacity to

consume and store high levels of nitrogen in nitrogensaturated watersheds. Such species can be favored, thus

increasing site nitrogen retention and reducing export.

Greater understanding is needed of the mechanisms and

capacities for nitrogen retention in various soils and

Watershed Processes—Fluxes of Water, Dissolved Constituents, and Sediment
39

ecosystems (Fenn and others 1998). Key indicators of

ecosystem nitrogen status need to be tested and implemented in monitoring networks in order to more fully

identify sites impacted by excess available nitrogen in the

ecosystem.

Key Points

1. The hydrologic cycle is highly coupled, so modifications

of one part of the system are likely to affect other parts

that may be far removed in time and space. It is important

to recognize the close coupling of surface water and

ground water systems and resources—failure to do so in

many past and present practices and policies has created

difficult problems in water allocation and environmental

protection.

2. Sediment production, transport, and storage should be

viewed as a complex system in which modification of

one part will affect other parts. On steep land, extreme

events commonly have profound, long-lasting effects on

sediment routing. Sediment impacts on drinking water

may not be strictly associated with present land management. Impacts may be partly attributed to land uses and

events that occurred many years previously.

3. For significant Federal projects, NEPA requires analysis

of the cumulative watershed effects, the aggregate

consequences of multiple land-use activities within a

watershed. Watershed effects can be addressed through

several complementary approaches. Watershed analysis

can provide broad, historical context for evaluating

potential cumulative effects of proposed land-use

activities. Thoughtful reviews of the issue (Reid 1993,

1998) describe prospects and pitfalls in addressing

cumulative watershed effects.

4. Watersheds in areas influenced by high atmospheric

nitrogen pollution from high population urban zones,

industrial areas, or in areas of mixed forest and intensive

irrigated and nitrogen-fertilized agricultural areas are at

risk of degraded water quality from nitrate concentrations

in surface and subsurface runoff. Equally important risk

factors include steep slopes and coarse-textured, shallow

soils; mature forests or vegetation with low-nitrogen

demand; high accumulation of nitrogen in organic matter;

rapid nitrogen cycling rates in soil and vegetation; and

the abundance of vegetation with high nitrogen fixation

rates, e.g., alder (Alnus spp.).

5. Management strategies for nitrogen-saturated watersheds

have not been adequately tested, but ecological principles

and past studies of nitrate runoff responses to silvicultural

treatments suggest reasonable strategies for reducing

nitrate runoff. The basic strategies include: (1) increasing

plant nitrogen demand, (2) reducing the amount of

nitrogen in the ecosystem, or (3) enhancing gaseous

losses of nitrogen (denitrification)—usually from the

riparian zone. These objectives may be accomplished by:

(1) stimulating forest production through thinning,

planting, harvest and regeneration, and fertilizing with

limiting nutrients other than nitrogen; (2) removing

nitrogen through prescribed burning and whole-tree

harvesting; and (3) discouraging transport by maintaining

effective vegetation buffer strips in the riparian zone.

Field studies are needed to test the effectiveness of these

approaches in a variety of ecosystem types and conditions. Most management options to mitigate nitrogensaturation effects are probably not applicable in wilderness areas.

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Chapter 3
42

1Economist, USDA Forest Service, Rocky Mountain Research Station, Fort

Collins, CO.

Chapter 4

Economic Issues for Watersheds Supplying Drinking Water

Thomas C. Brown1

Introduction

No other resource serves as many purposes as water. It is

widely used in industry, in electric energy production, in

farming and ranching, and, of course, by households for

drinking, washing, and gardening. Water is essential to the

health of ecological systems, supports numerous forms of

recreation, and provides important amenity values. In

addition, water is valuable in flushing and treating wastes,

both from contained sites such as industrial plants, commercial establishments, and houses, and from land areas such as

lawns, farms, and forests. Unfortunately, the processing of

wastes often leaves water unsuitable for other uses without

restoration of purity.

Water is essential to the viability of forests, farms, pastures,

and other land areas, but, as it runs off, water carries soil

from the land. Excess soil reaching streams impairs fish

habitat, accumulates in reservoirs and other water management facilities, and increases costs of water treatment. In

addition, pesticides, nutrients, and other contaminants

attached to soil particles often leave the site.

Water supply and water quality are thus integrally linked.

Most water users—whether they be boaters, farmers,

industries, or households—are affected by the water’s

quality and in turn affect quality of the water that others use.

These interdependencies make both water treatment and

watershed management essential.

Sources of water pollution are usually grouped into point

and nonpoint categories. Point sources, which emit from

pipes or canals, include municipal wastewater treatment

plants and industrial facilities. Nonpoint sources, which are

diffuse and difficult to monitor, include runoff from farms,

pastures, forests, cities, and highways, as well as rural septic

systems and landfills. Watershed management is, in large

part, the management of nonpoint sources of water

pollution.

Nonpoint sources have long been recognized as the primary

causes of some types of water pollution. For example,

Gianessi and Peskin (1981) estimated that in the 1970’s, 98

percent of the total suspended solids, over 85 percent of the

phosphorus and nitrogen, and 57 percent of the 5-day

biochemical oxygen demand in U.S. waters were attributable to nonpoint sources. For 1986, the U.S. Environmental

Protection Agency (EPA) (1987) reported that nonpointsource pollution was the cause of 65 percent of the waterquality-impaired stream miles and 76 percent of the

impaired lake acres. The most recent EPA water-quality

inventory, for 1996, reports a similar finding and shows that

although agriculture is by far the largest nonpoint source of

water pollution in the United States, forestry and other

activities are important sources in some areas (U.S. EPA

1998).

Since the 1972 Clean Water Act was passed, some progress

has been made in improving the Nation’s water quality. For

example, Lettenmaier and others (1991) examined trends

from 1978 to 1987 at 403 stations in the U.S. Geological

Survey’s National Stream Quality Accounting Network and

found significantly more stations with decreases than

increases in pathogens, oxygen deficit, phosphorus, and

some heavy metals. However, increases outnumbered

decreases for total nitrogen, and suspended sediment had

remained largely unchanged. In general, the successes are

associated with point-source controls and the lack of success

with nonpoint sources. Such findings suggest that the

Nation’s water-quality goals will not be met without

increased emphasis on nonpoint-source pollution.

The provision of high-quality drinking water is affected by a

host of natural events and human activities occurring on

watersheds. The natural events include extreme precipitation

events, forest fire, landslides, and transmission of pathogens

by wild animals, e.g., Giardia spp. The human activities

include mining, agricultural tillage, industrial production,

timber harvest, livestock grazing, automobile use, road

construction and maintenance (including deicing), and use

of fertilizers and pesticides (whether in agriculture, forestry,

range management, or by homeowners). The interactions

among these factors, and the unpredictable nature of some

factors, make water-quality protection a challenging task.
43

The costs of water-quality control in the United States are

substantial and rising. In 1985, households obtaining their

water from municipal systems spent 0.6 percent of their

income for water, plus an additional 0.4 percent for wastewater treatment (Singh and others 1988). These costs were

expected to increase by about 30 percent in response to

stricter standards implemented since the late 1980’s. A

recent EPA survey indicates that community water systems

in the United States will need to invest $138 billion over the

next 20 years (Hertzler and Davies 1997). Additional

expenditures will be necessary by industry, agriculture, and

other sectors to protect water quality. These costs highlight

the importance of considering the economics of water

quality.

Perhaps the most fundamental economic question regarding

drinking water quality is whether the benefits of drinking

water standards exceed the costs. The benefits consist of

averted losses of two general kinds. First, a water-quality

standard can avert losses from drinking unclean water,

including human health losses and associated health care

costs. Second, where meeting the standard involves controlling upstream sources of pollution, the benefits also include

averted losses between the pollution source and the drinking

water diversion, including fish population losses, costs of

removing sediment from canals and reservoirs, and decreases in recreation quality and use. The costs to be

compared with such benefits include those at water treatment plants or by rural households that must treat their own

water, and costs of controlling pollution emissions upstream

of the drinking water diversion. Potential upstream pollution

control costs include, for example, crop losses from decreased pesticide use; costs of controlling erosion from

fields, forests, and roads; reduced beef production associated with fencing cattle out of riparian areas; and costs at

upstream wastewater treatment plants.2

Despite serious efforts to estimate the benefits of drinking

water standards and other water-quality controls (Freeman

1982, 1993), the estimates remain rough. Because of

imprecise benefit estimates and reluctance to compromise

on the safety of public drinking water, drinking water

standards are often set without definitive economic analysis.

Although benefit-cost comparison of drinking water

standards remains an important issue, a more limited—

though still challenging—role for economics is perhaps of

more immediate relevance. That role and the focus of this

chapter is helping to determine how the standards, once set,

should be met.

To avoid waste of resources, standards should be met

efficiently. A drinking water standard may be met solely by

treating existing water prior to use, or by a combination of

water treatment at points of use and pollution control

upstream where the water-quality problems originate.

Because pollution may occur at various points in the

watershed, corrective action may involve many different

costs. And because the costs of alternative actions can differ

considerably, opportunities for inefficiencies (or conversely

for cost savings) abound.

A related economic issue is the equity of options for

implementing the efficient cost allocation.3 Expecting each

actor to bear the cost of any change required to minimize the

total cost of reaching the downstream water-quality standard

may unfairly allocate the costs. If so, options for cost

sharing, including the use of economic subsidies, should be

explored. These two issues, efficiency and equity, are

addressed below.4

Cost Minimization

Concerns about drinking water quality involve a relation

between upstream emitters of a pollutant and downstream

receptors who must treat the water before it can be safely

used. An emitter is any pollution source, such as a forest

area, a farm, or an urban wastewater treatment plant. A

receptor is any drinking water provider or rural domestic

user not served by a water provider.

The goal of a drinking water provider at a given use point j

is to reduce the concentration of a pollutant in water

delivered to users (Xd) to a level at or below the standard

(Xs):

(1)

For a water provider, achieving the desired water quality is a

function of the concentration of a pollutant at the reception

point (Xr) and of the reduction in that concentration by

treatment (T) before the water is used:

(2)

2

 For examples of such costs, see Easter (1993) on effects of reduced

herbicide use; Chang and others (1994), Lyon and Farrow (1995), and

Young and others (1991) on agricultural erosion costs; and Binkley and

Brown (1993b) on erosion control costs in forestry.

3

 Another economic issue, which under conditions of full employment is

largely a matter of equity or distribution, is the economic impact of

pollution in terms of jobs and income. These economic impacts are not

discussed herein.

4

 Several excellent books cover the topics summarized here, including

Freeman and others (1973), Kneese and Bower (1968), and Tietenberg

(1988).

Xdj ≤ Xs

Xdj Xr = −j Tj

Chapter 4
44

The receptor must react to Xr, increasing the level of

treatment to compensate for an increase in Xr.

Pollutant concentrations at the reception point are the result

of many upstream management actions and natural events.

For example, in figure 4.1, the city’s water treatment plant

receives pollution from the forest, the recreation area, the

upstream town’s wastewater treatment plant and storm

runoff, septic systems of rural households, and farms.

Pollutants from land areas such as forests and farms may

result from both natural (sometimes called background) and

management-caused emissions. In addition, upstream

consumptive use, such as by farms, towns, and transbasin

diversions, can increase the concentration of pollutants

reaching the receptor, and natural processing of pollutants

occurring in the stream or the adjacent alluvium decreases

the concentration.

Therefore, the concentration of a pollutant at reception point

j (Xrj

) is a function of the emissions (e) of each upstream

source (i), the transfer coefficients between each source and

the receptor (αij), and the amount of streamflow at the

reception point (Qj):

 (3)

The streamflow amount is equal to the natural, i.e., virgin,

flow minus upstream consumptive use resulting from each

upstream diversion. The transfer coefficient α varies from 0

to 1 and reflects the water treatment that naturally occurs

between the emission and the receptor, plus any removal of

pollutants by diversions. For degradable pollutants, naturally

occurring treatment increases (causing α to decrease) with

distance, all else equal; for nondegradable pollutants α

approaches 1. Removal of pollutants by diversions is most

common with transbasin diversions; other diversions may

temporarily remove some pollutants, but they often eventually return to the stream with return flows.

If the water body is a lake rather than a stream, equations (1)

through (3) apply, but Q is storage rather than flow; all

receptors on the lake are potentially affected by all emitters,

and α for a given pollutant will not differ among emissions

to the extent that mixing occurs.

The economic task is to determine the most cost-effective

way to reach the goal characterized in equation (1). Pollution can be controlled at its source or removed at the point

of reception and can be lessened by dilution. Hence,

upstream emitters, upstream consumptive users, and

downstream receptors are all candidates for actions to help

meet the drinking water standard.5 Each actor has a cost of

reducing the concentration of a pollutant to the required

level. Ideally, from an efficiency point of view, control

efforts would occur at the most cost-effective points.

The cost at upstream pollution source i (Ceij) depends on the

reduction in profit or benefit caused by reducing the

concentration of the emission that reaches receptor j. The

cost at upstream consumptive use point k (Cukj) is the

reduction in profit or benefit caused by reducing consumptive use so that more water reaches receptor j, thereby

reducing the concentration of pollutants.6 The cost at the

point of reception (Ctj) is the cost of water treatment prior to

use. The objective for use point j is to minimize the total

cost of meeting the standard (Cj) where:

(4)

Figure 4.1—Hypothetical river basin.

5

 Emitters may also be diverters, and, therefore, potentially consumptive

users, e.g., agricultural irrigators and cities, but not all emitters are

diverters, e.g., forests, and not all diverters are emitters, e.g., transbasin

water diversions.

6

 The cost for an upstream consumptive user is more complicated than

expressed in equation (4) when the water pollutants removed from the

stream with the diversion do not all return to the stream in return flow. In

this case, the reduction-in-treatment cost at the downstream drinking water

treatment plant caused by the upstream removal of pollutants with the

diversion must be subtracted from the increase in cost at the treatment plant

caused by decreases in streamflow that occurs with the consumptive use of

the diversion. Obviously, when the diverted pollutants do not return to the

stream, the more polluted is the diverted water, the lesser is the cost

imposed by the consumptive use on the downstream drinking water

treatment plant.

Forest

City

Town

Farm

Transbasin

diversion

Farm

Recreation

Rural

houses

Xrj ei

i

= ∑ αij Qi

Cj Ceij

i

Cukj Ctj

k

=++ ∑ ∑

Economic Issues for Watersheds Supplying Drinking Water
45

The costs of each entity are a function of prices of inputs

and outputs and of the entities’ production functions and

how their production actions affect Xdj. Of course, if a

watershed has more than one reception point, the overall

cost efficiency goal is to minimize the sum of the various Cj

.

Opportunities for Cost Savings

To find opportunities for cost savings, we must understand

the costs of emitters, consumptive users, and receptors. In

the short run, most of the cost of a water treatment physical

plant is fixed, and only variable costs (for labor, materials,

and supplies) change with changes in the concentration of

pollutants entering the plant or with the volume of water

treated. Similarly, in the short run, emitters’ and consumptive users’ facilities or equipment, such as timber harvest

machinery, road designs, livestock fences, agricultural

irrigation machinery, homeowners’ septic systems, and canal

sizes, are fixed. However, in the long run, fixed costs

change to permit expansion of existing treatment facilities,

introduction of new machinery, fencing, canal lining, etc.

Thus, flexibility to adapt to changing levels of pollutant

concentration, changing drinking water demands, or

changing water-quality standards is much greater in the long

run.

Short-Run Costs

Marginal cost curves, showing the change in cost with a

change in some measure of output, can be estimated for the

short or long run. Consider first the short-run cost curve of

an upstream pollution source such as a forest road, expressed as a function of pollutant concentration (fig. 4.2).7

If no effort is made to control emissions (in this case,

sediments), the concentration of the pollutant reaching a

water-use reception point is Xr’ and, of course, the emitter’s

marginal cost of control is $0. Initial reductions in the

concentration of the pollutant reaching the reception point

are likely to be relatively inexpensive, perhaps brought

about by cleaning culverts and drainage ditches. However,

further reductions in the pollutant concentration are likely to

be progressively more expensive, as indicated by the

increasing emitter marginal cost in figure 4.2.8 Reducing the

concentration to zero may be quite expensive, and could

require closing the road altogether.9

Now consider the short-run marginal cost curve of a

downstream water treatment plant. This curve (the receptor

treatment cost curve in figure 4.2)10 also is likely to rise as

the pollutant concentration level is lowered because even

lower concentrations are more and more costly to achieve.11

However, the receptor’s marginal cost is unlikely to drop to

zero at a high level of concentration, as does the emitter’s

cost curve because of the need to maintain the labor and

Figure 4.2—Efficient allocation of cost of meeting drinking water quality

standard, with one emitter and one receptor.

7

 The appropriate marginal cost curve for an emitter takes account of the

natural assimilative capacity of the environment for the pollutant at issue [α

in equation (3)]; it depicts the marginal cost of reducing the pollutant load

at the point of reception (Xr), not at the point of emission. If two emitters

yield identical amounts of pollution but have different transfer coefficients,

their effective marginal cost curves from the standpoint of meeting the

drinking water-quality standard, are different.

8

 Marginal cost curves are typically drawn with movement to the right

along the horizontal axis indicating increasing producer effort, so that the

marginal cost curve has a positive slope. Because the horizontal axis in

figure 4.2 is concentration of a pollutant, producer effort increases to the

left and, thus, the marginal cost curve has a negative slope.

9 The marginal cost curve of a consumptive user who is not an emitter, such

as a transbasin diversion of pristine water, is likely to be similar in shape to

that of the emitter shown in figure 4.2. That is, initial reductions in

diversion are likely to be inexpensive, especially where water use is

subsidized, as is much irritation in the West. However, further reductions

will only be possible at increasing costs.

10 This curve assumes a given volume of water treated to the concentration

level indicated on the horizontal axis. The entire curve shifts up as water

volume increases. A treatment plant’s marginal cost curve could also be

expressed as a function of volume of water treated assuming a constant

level of treatment, i.e., a constant level of concentration reduction, per unit

of water volume. The marginal cost curve in this case would have a positive

slope, and would shift vertically with changes in the treatment level. A

three-dimensional graph could, of course, show marginal cost as a function

of both volume of water treated and treatment level.

11 Moore and McCarl (1987) provide data for plotting a water treatment

plant marginal cost curve. They estimated the marginal costs of removing

sediment at a municipal water treatment plant in Corvallis, OR. The

principal costs modeled were for alum, lime, and sediment disposal. Over a

wide range in sediment concentration, marginal cost increased only slightly

as sediment concentration decreased, but as the concentration approached

zero the marginal cost abruptly increased.

Xs Xr* Xr1

Emitter

control cost

Receptor

treatment cost

Emitter's

total cost

Receptor's

total cost

Concentration of pollutant at point of reception

Marginal cost ($)

Chapter 4
46

Economic Issues for Watersheds Supplying Drinking Water

materials necessary to meet a water-quality threat at all

times. Even if the water-quality standard was set as low as

Xr’, the provider would still need to maintain the daily

capability of handling water withdrawals with pollutant

concentrations that exceeded Xr’. Thus, as shown in figure

4.2, the receptor’s marginal cost curve flattens out to the

right but remains above the horizontal axis.

Assuming a single emitter and single receptor represented

by the two short-run marginal cost curves of figure 4.2, and

a water-quality goal no greater than Xs, the efficient

allocation of treatment and control costs is indicated by the

intersection of the two cost curves at a concentration of Xr*.

To the right of Xr*, the emitter’s marginal cost is lower than

the receptor’s, whereas to the left the reverse is true.

Requiring the emitter to reduce the concentration at the

point of reception below Xr* costs the emitter more than it

saves the receptor, and not requiring the emitter to reduce

the concentration to at least Xr* costs the receptor more than

it saves the emitter.

The sum of the total costs, equivalent to Cj

 in equation (4),

is minimized by finding the equimarginal point along the

relevant marginal cost curves. Total cost is equal to the area

below the relevant marginal cost curve. Assuming an

efficient distribution of costs as in figure 4.2, the emitter’s

total cost is represented by the area below the emitter’s

marginal cost curve to the right of Xr*, and the receptor’s

total cost is the area below the receptor’s marginal cost

curve to the left of Xr* and right of Xs.

Although the receptor’s marginal cost curve will always be

above the emitter’s at concentration level Xr’ (fig. 4.2), the

emitter’s marginal cost curve will not necessarily rise above

the receptor’s as the concentration level is reduced. If the

emitter’s marginal cost curve remains below the receptor’s

curve at all concentration levels, costs are minimized by

focusing all pollution control efforts on the emission source.

The precise placement of the emitter’s and receptor’s cost

curves may be difficult to determine. And the marginal costs

of the two entities could be quite similar over some range in

concentration, further complicating determination of the

equimarginal point. However, in some cases the opportunities for cost savings will be obvious; it is these cases where

cost control efforts should initially focus. For example,

consider costs of phosphorus reduction for agriculture

versus municipal treatment plants. Schleich and others

(1996) report average costs to reduce a kilogram of phosphorus of $26 using onsite pollution control practices in

agriculture and $169 at municipal treatment plants (1990

dollars).

Long-Run Costs

Often, cost minimization involves long-run decisions. Longrun cost curves of water treatment plants depict how costs

change as plant capacity increases to handle a given

pollutant. Most such curves have focused on changing water

volumes; they typically show economies of scale, with

considerable decreases in average costs as plant size

increases, along with decreasing or relatively constant longrun marginal cost curves. For example, figure 4.3 shows

construction cost for a pressure filtration plant as estimated

by Gumerman and others (1978), expressed in 1978

dollars.12

When comparing treatment plant costs with costs of

controlling pollution at its source, the most relevant issue is

pollutant concentration rather than water volume. The

relevant long-run marginal cost at the treatment plant may

be the cost of adding or altering, not simply expanding, a

treatment capability to deal with increased pollution

concentrations (rising Xr) or tightening of water-quality

standards (lowering of Xs). For example, the oocysts of the

protozoan Cryptosporidium spp. are not inactivated by

chlorine at dosages that are feasible in drinking water

treatment. If these oocysts must be removed at a treatment

plant that has relied on chlorine to control pathogens, new

processes, such as filtration or use of ozone, will be necessary. In such a case, the long-run marginal cost curve as a

function of concentration in the received water rises

abruptly at a concentration equal to the water-quality goal

(Xs) of the drinking water provider, as in figure 4.4. As

12 Computer models now exist for estimating treatment costs for a variety

of water treatment processes; see Raucher and others (1995) for summaries

of several such models.

Figure 4.3—Construction cost of pressure filtration plant with an

infiltration rate of 2 gallons per minute per square foot (Gummerman

and others 1978).

0

10

5

50,000 100,000

Average cost per gallon per day

Marginal cost per gallon per day

Gallons per day

Cost ($)
47

discussed later regarding New York City, upstream pollution

control may help avoid such upward jumps in marginal cost.

Complexity

Although straightforward in concept, minimizing cost from

the nonpoint-source-pollution context is extremely difficult

in practice, principally because of the complexity of the

physical processes involved. Numerous factors complicate

the cost minimization. First and foremost, nonpoint-source

pollution, by its very nature, is difficult to monitor at its

source, especially on a continuous and widespread basis.

Downstream water quality may be assessed, but linking that

water quality to upstream events and locations is inexact at

best. Even in the case of sediment and other natural pollutants, it is often difficult to separate user emissions from

background levels.

Six additional factors further complicate assessment and

minimization of the costs of meeting drinking water goals:

• There may be numerous pollution sources and numerous

points of consumptive use, so computing the minimum

cost for a given receptor may require estimating many

different costs.

• A basin is likely to have numerous drinking water

reception points.

• Each emitter and receptor must be concerned with

numerous, different pollutants and treatment or control of

one pollutant may affect other pollutants. For example,

treatment for Giardia may remove other pathogens, and

erosion control will reduce transmission of pesticide

residues.

• Each emitter, consumptive user, and diverter may have

several options for lowering Xr. For example, a forest

manager may lower stream sediment loads by more

careful placement of skid trails, improved engineering of

roads, and avoiding harvest near streams.

• Xr is stochastic, depending on unpredictable (and perhaps

highly intermittent) weather events and uncertain actions

of upstream landowners.13

• Uses beyond drinking water, such as fish habitat, recreational swimming, and industrial water use, are affected

by the quality of the water in the stream or lake. If

society’s cost efficiency objective is to minimize the total

cost of reaching its various water-quality goals in a

watershed, pollution control decisions must take all water

uses into account. The resulting cost minimization will

involve a mix of instream and drinking water-quality

standards.

The difficulty of measuring many of the components of a

watershed’s pollution control cost minimization problem,

and the random nature of nonpoint-source pollution,

contribute to a high level of uncertainty about the effects of

upstream nonpoint-source pollution and efforts to control it

on downstream pollution levels and treatment costs (Shortle

1987). Because of this uncertainty, it is often difficult to

know just what to do and where to do it to minimize costs of

meeting water-quality goals. Water-quality control in this

context must, therefore, be iterative, localized, continuous,

and long term—iterative because the parties involved will

learn by doing, localized because the solutions will be

highly site-specific, continuous because vigilant monitoring

is necessary to assess compliance and fine tune the control

effort, and long term because nonpoint-source pollution

depends on extreme and, thus, infrequent weather events.

This complexity should not unduly detract, however, from

the central point that opportunities for cost savings may

exist, as seen in the next section.

13 Forest lands demonstrate this point. Although not generally a significant

source of nonpoint-source pollution (Binkley and Brown 1993a), soil loss

from such lands can be substantial in the wake of severe weather events.

Erosion can be particularly serious if severe weather happens to coincide

with activities that temporarily expose soil, such as forest fire, timber

harvest, and road construction. Also, protecting the forest from harvest and

associated roads is not necessarily the best policy for protecting water

quality, because natural fuel buildup may lead to more devastating fires

and, thus, to greater eventual soil loss. See Brown and others (1993) for

more on the policy and economics of nonpoint-source pollution control in

forest areas.

Figure 4.4—Long-term marginal cost as a function of pollutant concentration.

Xs

Concentration of a pollutant at reception point

Marginal cost ($)

Chapter 4
48

Cost Savings from Targeting Upstream

Control Efforts

Several studies have estimated the cost savings obtained by

replacing so-called command and control strategies of

pollution control, which emphasize uniform controls across

all emitters or all subwatersheds, with careful targeting of

upstream control efforts. An early study (Johnson 1967)

examined dissolved oxygen levels in the Delaware River

Basin using a model that identified the major pollution

sources and tracked pollution levels. The study compared a

uniform percentage reduction in oxygen-demanding wastes

from all polluters with three more cost-effective distributions of control efforts. Depending on the stringency of the

dissolved oxygen goal, the uniform control strategy was

from 1.4 to 3.1 times as costly as the most inexpensive

strategy of carefully targeted control efforts.14

Schleich and others (1996) studied the costs associated with

reducing phosphorus levels in the Fox-Wolf River Basin by

50 percent. They compared costs of meeting the target in

each of 41 subwatersheds with meeting the target at the

river’s mouth in Green Bay. Municipal, construction, and

agricultural emissions were modeled. Meeting the goal at

each subwatershed was 4.5 times more expensive than the

basin-wide strategy of only meeting the goal in Green Bay.

With the basin-wide strategy, only 19 sources (18 of them

agricultural) are selected for phosphorus reduction. The

primary cost savings occur from not forcing watersheds with

already low levels of phosphorus emissions (usually those

without major agricultural sources) to participate in the

proportional reduction scheme; further savings accrue from

consideration of loading factor differences among

subwatersheds.

Other studies have focused on the command and control

strategy of requiring each emitter to reduce pollution

loading to a specified level. Although more sensible than

proportional reductions, this strategy also fails to minimize

costs because it ignores differences in emitters’ control

costs. Studies involving largely point-source pollution have

repeatedly shown that savings can be achieved by using a

control strategy that allows differential amounts of control

as long as the downstream or ambient goal is reached.

Tietenberg (1985) and Anderson and others (1997) summarize these studies.

Most economic studies of nonpoint-source pollution have

dealt with agriculture. Several have demonstrated how costs

of reaching downstream water-quality goals are minimized

by carefully selecting pollution control locations and levels.

For example, studies of soil loss from a 1,064-acre watershed in Illinois (Braden and others 1989) and a 11,400-acre

watershed in Minnesota (Kozloff and others 1992) found

significant cost savings in meeting downstream waterquality goals from taking into account the farm-specific

costs of reducing emissions as well as loading factor

differences. In the Illinois study, careful targeting allowed

the area requiring changes in management to be reduced by

roughly 80 percent; targeted changes were concentrated near

streams and involved mainly restrictions on crop rotation

and tillage. In the Minnesota study, farmers’ control costs

were reduced by one-half or more when control efforts were

carefully targeted.

Bringing About an Efficient Cost Allocation

Much of the economic writing on pollution (such as Baumol

and Oates 1975, Freeman 1990, Freeman and others 1973,

Kneese and Bower 1979, Tietenberg 1985) focuses on how

to structure economic incentives to efficiently meet pollution control objectives. The theory for structuring economic

incentives was developed primarily for point-source

pollution, for which efficient mechanisms like emission

taxes or subsidies and tradable permits can work well.

Several European countries and more recently the United

States as well have made much progress in using these

mechanisms to efficiently control point sources of air and

water pollution. The mechanisms have not, however, been

easily adapted to the control of nonpoint-source pollution.

The principal problem in designing an economic incentive

mechanism for control of nonpoint-source water pollution is

that nonpoint-source emissions are stochastic and difficult to

measure at their point of origination. Without linking

pollution to specific land parcels, there is no way to accurately charge a tax, offer a subsidy, or trade a permit. A way

around the measurement problem is to approximate measurement using a predictive model. However, the complexities of soil and pollutant movement, and the resultant errors

in prediction, have hindered development of appropriate

models. Because of this lack of measurement or modeling

precision, plus a political unwillingness to force landowners

to comply, the major efforts at nonpoint-source pollution

have focused on education about and voluntary adoption of

pollution control practices, plus government incentives to

retire highly erosive land from agriculture. Although the

incentives have had some success (Freeman 1990, Ribaudo

1989), it is claimed that education and most other

nonregulatory approaches have failed to provide sufficient

motivation for major changes (Adler 1992). 14 Tietenberg (1985) summarizes two additional biological oxygen demand

studies with similar results to the Delaware River Study.

Economic Issues for Watersheds Supplying Drinking Water
49

Although nonpoint-source emissions cannot be as effectively taxed or traded as point-source emissions, there

remain considerable economic incentives for downstream

drinking water providers to negotiate with upstream

polluters because the downstream providers must ultimately

meet drinking water standards in order to protect human

health. In the absence of enforceable regulations requiring

upstream polluters to alter their behavior, such negotiations

are likely to take the form of the downstream drinking water

providers paying the upstream polluters to follow practices

that are thought to reduce emissions. These negotiations

have been called point and nonpoint-source or both pollution trading, but essentially they are a subsidy scheme

(Malik and others 1994).

A problem with subsidies is that polluters have an incentive

to cease voluntary control practices, or even to adopt

polluting practices, in order to become more attractive

candidates for a subsidy (Baumol and Oates 1975, Malik

and others 1994). For the subsidy scheme to work, therefore,

it may be necessary to impose some watershed-wide

minimum pollution control practices that are sufficiently

fundamental and inexpensive as to be politically feasible.

The subsidies would then fund additional nonpoint-source

pollution control efforts, building on the baseline established

by the required practices. State efforts to specify and reach

instream water-quality standards, pursuant to the Clean

Water Act, may help provide this baseline.

Although one may argue that property owners should not

have to be paid to not pollute, subsidies may be more fair

and are often more politically feasible than additional landuse regulations. A recent agreement between New York City

and watershed landowners is a prime example of this

approach.

The New York City Agreement

The Catskill and Delaware watersheds, an area of roughly

1,600 square miles [4100 square kilometers (km2)], provide

90 percent of New York City’s water supply. Because of past

efforts at watershed protection, a series of city-owned

reservoirs that allows long detention times and flexibility in

meeting demands, and the low population density in the

watersheds, the city has avoided installing filtration for this

system (Ashendorff and others 1997).15 However, new

concerns about pathogens (specifically Giardia and

Cryptosporidium) and about economic growth in the

watershed have increased pressures for filtration, leading to

a 1997 agreement between the city and the EPA.

With the agreement, the city avoided, at least until the year

2002, the high cost of filtration, estimated at from $4 to $8

billion (Okun and others 1997). Instead, the city will invest

approximately $1.2 billion over the next few years in efforts

to protect the quality of the water entering the city’s water

treatment plants.16 Components of this investment include

the following:

• Upgrading the nine wastewater treatment plants that the

city operates for upstream communities.

• Rehabilitating and upgrading city-owned dams and water

supply facilities.

• Purchasing land and conservation easements in the

watershed.

• Funding various efforts of noncity entities, such as

inspection and rehabilitation of septic systems; improvements of sewer systems; better stormwater management;

environmental education; stream corridor protection; and

improved storage of sand, salt, and deicing materials.

• Paying farmers to follow best management practices.

• Enhanced monitoring.

In addition, the agreement places restrictions in the watershed on the siting of new wastewater treatment plants, the

operation of wastewater treatment plants, the construction of

new septic systems, and storage of petroleum products and

hazardous substances.17

Benefits and Difficulties of Localized Negotiation

A benefit of direct negotiations between downstream water

providers and upstream polluters is that it localizes control

efforts at the watershed level, where the parties involved

have the greatest knowledge of watershed and water-quality

conditions and the largest incentive to bring about a costeffective agreement.

Another benefit of local watershed-based agreements is that

they allow for participation of parties concerned with water

uses occurring between the upstream control point and the

downstream treatment plant. These uses might include fish

habitat, reservoir and canal use, and instream recreation.

Such parties would benefit from the agreement but are often

15 New York City is unusual in this sense. Over 90 percent of surface water

systems in the United States use filtration (Raucher and others 1995).

16 The State Government will contribute another $53 million to foster

partnership initiatives and the Federal Government will contribute up to

$105 million under the Safe Drinking Water Act Amendments of 1996.

17 For more on the agreement, see the September 1999 issue of “Water

Resources Impact” (volume 1, number 5) published by the American Water

Resources Association, and the following Web sites: http://www.state.ny.us/

watershed and http://www.epa.gov/region02/water/nycshed.

Chapter 4
50

too poorly funded to initiate the process and may be enticed

to participate in an agreement initiated by the drinking water

provider.18

The efficacy of the subsidy approach hinges on what

economists call transaction costs, meaning the costs of

gathering necessary information, bringing the parties

together, negotiating the details, and monitoring compliance

with the agreement. Transaction costs are lower and, thus,

success is more likely, where the numbers of significant

emitters and of large, downstream users are small (Easter

1993).19

Conclusion

Minimizing the cost of meeting drinking water-quality goals

will require considering the full range of options for

controlling pollution at the source. However, the complexities and uncertainties of nonpoint-source pollution seriously

constrain efforts to utilize traditional economic incentives to

reach cost-efficiency goals. Nevertheless, real opportunities

exist for cost savings, which are most likely to be realized

by a combination of limited pollution control regulations to

provide a baseline of control and watershed-based negotiations that emphasize subsidies to encourage use of practices

thought to reduce nonpoint-source emissions. Initial efforts

will focus on the most obvious cost saving opportunities,

where the benefits of nonpoint-source pollution controls are

clear and the transaction costs are limited. Careful monitoring will then hopefully allow fine-tuning of existing control

efforts and addition of new ones where warranted.

Literature Cited

Adler, Robert W. 1992. Returning to the goals of the Clean Water Act.

Water Resources Update. 88: 2–30.

Anderson, Robert C.; Carlin, Alan; McGartland, Albert M.;

Weinberger, Jennifer B. 1997. Cost savings from the use of market

incentives for pollution control. In: Kosobud, Richard F.; Zimmerman,

Jennifer M.,eds. Market-based approaches to environmental policy:

regulatory innovations to the fore. New York: Van Nostrand Reinhold:

15–46.

Ashendorff, Arthur; Principe, Michael A.; Seeley, Anne [and others].

1997. Watershed protection for New York City’s supply. Journal of the

American Water Works Association. 89(3): 75–88.

Baumol, William J.; Oates, Wallace E. 1975. The theory of environmental policy. Englewood Cliffs, NJ: Prentice-Hall, Inc. 272 p.

Binkley, Dan; Brown, Thomas C. 1993a. Forest practices as nonpoint

sources of pollution in North America. Water Resources Bulletin. 29(5):

729–740.

Binkley, Dan; Brown, Thomas C. 1993b. Management impacts on water

quality of forests and rangelands. Gen. Tech. Rep. RM–239. Fort

Collins, CO: U.S. Department of Agriculture, Forest Service, Rocky

Mountain Forest and Range Experiment Station. 115 p.

Braden, John B.; Johnson, Gary V.; Bouzaher, Aziz; Miltz, David.

1989. Optimal spatial management of agricultural pollution. American

Journal of Agricultural Economics. 71(2): 404–413.

Brown, Thomas C.; Brown, Douglas; Binkely, Dan. 1993. Laws and

programs for controlling nonpoint source pollution in forest areas. Water

Resources Bulletin. 29(1): 1–13.

Chang, Ching-Cheng; Atwood, Jay D.; Alt, Klaus; McCarl, Bruce A.

1994. Ecomonic impacts of erosion management measures in coastal

drainage basins. Journal of Soil and Water Conservation. 49(6):

606–611.

Easter, K. William. 1993. Differences in the transaction costs of strategies

to control agricultural offsite and undersite damages. In: Russell,

Clifford S.; Shogren, Jason F., eds. Theory, modeling, and experiences in

the management of nonpoint-source pollution. Boston: Kluwer

Academic Publishers: 37–69.

Freeman, A. Myrick, III. 1982. Air and water pollution control: a benefitcost assessment. New York: John Wiley. 186 p.

Freeman, A. Myrick, III. 1990. Water pollution policy. In: Portney, Paul

R.; Dower, Roger C.; Freeman, A. Myrick [and others], eds. Public

policies for environmental protection. Washington, DC: Resources for

the Future: 87–149.

Freeman, A. Myrick, III. 1993. The measurement of environmental and

resource values: theory and methods. Washington, DC: Resources for the

Future. 516 p.

Freeman, A. Myrick, III; Haveman, Robert H.; Kneese, Allen V. 1973.

The economics of environmental policy. New York: John Wiley. 184 p.

Gianessi, Leonard; Peskin, Henry M. 1981. Analysis of national water

pollution control policies: 2. agricultural sediment control. Water

Resources Research. 17(4): 804–821.

18 Moore and McCarl (1987) offer one example of the mix of potential

downstream cost savings obtainable by upstream pollution control. They

estimated that 77 percent of the downstream costs of erosion in Oregon’s

Willamette Valley were attributable to road maintenance (mainly for ditch

and culvert cleaning), 18 percent were incurred at water treatment plants,

and the remaining 5 percent were incurred for river dredging to maintain

navigation. Data were insufficient to include costs related to fish habitat or

flooding.

19 The design of pollution control incentives in the context of the

complexity and uncertainty inherent with nonpoint-source pollution is

discussed in depth by, among others, Segerson, Shortle, and their

colleagues (Segerson 1988, 1990; Shortle and Abler 1997; Shortle and

Dunn 1986).

Economic Issues for Watersheds Supplying Drinking Water
51

Gumerman, Robert C.; Culp, Russell L.; Hansen, Sigurd. 1978.

Estimating costs for water treatment as a function of size and treatment

plant efficiency. EPA–600/2–78–182. Cincinnati: U.S. Environmental

Protection Agency, Municipal Environmental Research Laboratory.

282 p.

Hertzler, Patricia C.; Davies, Clive. 1997. The cost of infrastructure

needs. Journal of the American Water Works Association. 89(3): 55–61.

Johnson, Edwin L. 1967. A study in the economics of water quality

management. Water Resources Research. 3(2): 291–305.

Kneese, Allen V.; Bower, Blair T. 1968. Managing water quality:

economics, technology, and institutions. Washington, DC: Resources for

the Future. 328 p.

Kneese, Allen V.; Bower, Blair T. 1979. Environmental quality and

residuals management. Baltimore: Johns Hopkins University Press.

337 p.

Kozloff, Keith; Taff, Steven J.; Wang, Yingmin. 1992. Microtargeting the

acquisition of cropping rights to reduce nonpoint source water pollution.

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Lettenmaier, Dennis; Hooper, Eric R.; Wagoner, Colin; Faris, Kathleen

B. 1991. Trends in stream quality in the continental United States,

1978–1987. Water Resources Research. 27(3): 327–339.

Lyon, Randolph; Farrow, Scott. 1995. An economic analysis of Clean

Water Act issues. Water Resources Research. 31(1): 213–223.

Malik, Arun S.; Larson, Bruce A.; Ribaudo, Marc. 1994. Economic

incentives for agricultural nonpoint source pollution control. Water

Resources Bulletin. 30(3): 471–479.

Moore, Walter B.; McCarl, Bruce A. 1987. Off-site costs of soil erosion:

a case study in the Williamette Valley. Western Journal of Agricultural

Economics. 12(1): 42–49.

Okun, Daniel A.; Craun, Gunther F.; Edzwald, James K. [and others].

1997. New York City: to filter or not to filter. Journal of the American

Water Works Association. 89(3): 62–74.

Raucher, Robert S.; Castillo, Eloise Trabka; Dixon, Ann [and others].

1995. Estimating the cost of compliance with drinking water standards:

a user’s guide. Denver: American Water Works Association Research

Foundation. 380 p.

Ribaudo, Marc O. 1989. Water quality benefits from the conservation

reserve program. Agric. Econ. Rep. 606. Washington, DC: U.S.

Department of Agriculture, Economic Research Service. 30 p.

Schleich, Joachim; White, David; Stephenson, Kurt. 1996. Cost

implications in achieving alternative water quality targets. Water

Resources Research. 32(9): 2879–2884.

Segerson, Kathleen. 1988. Uncertainty and incentives for nonpoint

pollution control. Journal of Environmental Economics and Management. 15: 87–98.

Segerson, Kathleen. 1990. Incentive policies for control of agricultural

water pollution. In: Braden, John B.; Lovejoy, Stephen B., eds.

Agriculture and water quality: international perspectives. Boulder, CO:

Lynne Rienner Publishers: 39–61.

Shortle, James S. 1987. Allocative implications of comparisons between

the marginal costs of point and nonpoint source pollution abatement.

Northeastern Journal of Agricultural and Resource Economics. 16:

17–23.

Shortle, James; Abler, David G. 1997. Nonpoint pollution. In: Folmer,

Henk; Tietenberg, Tom, eds. The international yearbook of environmental and resource economics 1997/1998: a survey of current issues.

Cheltenham, UK: Edward Elgar Publishing Limited: 114–155.

Shortle, James S.; Dunn, James W. 1986. The relative efficiency of

agricultural source water pollution control policies. American Journal

of Agricultural Economics. 68(3): 668–677.

Singh, Jasbinder; Stein, Raffael; Chandra, Sanjay; Snyder, Brett. 1988.

The municipal sector study. EPA 230–09/88–038. Washington, DC: U.S.

Environmental Protection Agency. 166 p.

Tietenberg, T.H. 1985. Emissions trading: an exercise in reforming

pollution policy. Washington, DC: Resources for the Future. 222 p.

Tietenberg, Tom. 1988. Environmental and natural resource economics.

Glenview, IL: Scott, Foresman, and Company. 559 p.

U.S. EPA. 1987. National water quality inventory: 1986 report to Congress.

EPA–440/4–87–008. Washington, DC: U.S. Environmental Protection

Agency. 141 p.

U.S. EPA. 1998. National water quality inventory: 1996 report to Congress.

EPA–841–R–97–008. Washington, DC: U.S. Environmental Protection

Agency. 521 p.

Young, Douglas L.; Walker, David J.; Kanjo, Paul L. 1991. Cost

effectiveness and equity aspects of soil conservation programs in a

highly erodible region. American Journal of Agricultural Economics.

73(4): 1053–1062.

Chapter 4
53

Part II:

Effects of Recreation

and the Built Environment

on Water Quality

Canoeists at Ozark Landing, Buffalo National Wild and Scenic River,

Buffalo National River, Arkansas. Photo by Bill Lea
55

Chapter 5

Hydromodifications—Dams, Diversions, Return Flows,

and Other Alterations of Natural Water Flows

Stephen P. Glasser1

Introduction

The term hydromodification is commonly used to describe

all activities, which alter the natural flow of water. This

chapter addresses the effects of structures, such as dams,

headgates, reservoirs, canals, water wells, diversion ditches,

and flumes upon the quality of raw drinking water before it

arrives at the water treatment plant. It also includes a

discussion of the effects of land application of treated

sewage sludge, return flows, wetland modifications, and

reclaiming wastewater upon drinking water quality.

Issues and Risks

The U.S. Environmental Protection Agency (EPA) ranked

hydromodification as the third leading cause of waterquality impairment to rivers. Only agriculture and municipal

sewage treatment plants ranked higher (U.S. EPA 1995).

Nationwide, there are over 68,000 medium and large dams

built for hydropower, water supply, flood control, and other

purposes. The U.S. Geological Survey estimates the

cumulative storage capacity of these dams is almost 450

million acre-feet [550 billion meters (m3

)].2 The Bureau of

Reclamation manages about 600 dams and 53,000 miles

[85000 kilometers (km)] of canals in 17 Western States; the

Army Corps of Engineers has about 700 dams and accounts

for about one-third of all water in storage in the Nation

(Reetz and others 1998). There are about 2,350 dams with a

total storage capacity of about 55 million acre-feet (68

billion m3) on land administered by the U.S. Department of

Agriculture, Forest Service. Half are owned and operated by

the Agency mostly for recreation, fire protection, and fish or

wildlife needs. The others are owned and operated by other

Federal agencies, States, and private parties under specialuse authorizations, mostly for irrigation, recreation, and

water supply.3

There are also thousands of small dams in the United States

that were designed and built to store drinking water during

periods when inflows to the reservoir are greater than the

water removed from the reservoir. Some of these reservoirs

were built and are still operated solely to provide a reliable

water supply. Since the 1940’s, however, some of the

existing ones and almost all new reservoirs became multipurpose; that is, they serve recreation, irrigation, flood

control, and sometimes hydropower needs, while supplying

drinking water. Often these other purposes create waterquality problems for human health by altering water

temperature, sediment transport, biological oxygen demand,

chemical oxygen demand, total dissolved solids, and

streamflow. Related information on these problems can be

found in chapters 2 and 3.

The diversion and transport of water from one watershed to

another can result in physical, chemical, and microbiological contamination of the receiving waterbody and cause

channel erosion, sediment transport, and deposition in

reservoirs and channels. Subsequent dredging in large rivers

and reservoirs often accelerates downcutting of headwater

streams and destabilizes streambanks, even where stream

gradients are quite flat, such as in Mississippi.

Drainage of wetlands with ditches is a form of

hydromodification that can change water chemistry by

adding organic compounds, thereby affecting water treatment processes and costs. Application of treated sewage

sludge to forested land has been evaluated for its risk of

contaminating water supplies with pathogens and found to

be a low risk in most situations. Reclaiming sewage effluent

water for drinking water is done in other countries, but is not

yet commonplace in the United States.

Nearly all these hydromodifications are influenced by water

rights laws which vary considerably from State to State. In

most Western States, laws require water users to divert

water out of streams or rivers to obtain a State water right.

This removal often results in higher water temperatures,

lower oxygen levels, reduced sediment transport capacity,

and other water-quality problems in the remaining water

1

 Water Rights and Uses Program Manager, USDA Forest Service,

Washington, DC.

2

 Personal communication. 1999. Robert Hirsch, Chief Hydrologist, U.S.

Geological Survey, 807 National Center, Reston, VA 20192.

3 Personal communication. 1998. James Padgett, Chief Hydraulic Engineer,

USDA Forest Service, Washington, DC 20250.
56

(Getches and others 1991). The riparian water rights

doctrine used in most States east of the 100th meridian

generates water-quality problems because most of the water

is returned to the channel.

Findings

Hydromodifications can impact water quality via algae

blooms, trihalomethane production, sediment transport and

deposition, and changes in chemical, physical, and microbiological properties.

Effects of Dams and Impoundments

on Water Quality

The size and depth of impoundments and the residence time

of water within them can affect water quality chemically,

physically, and biologically. As water flows into a reservoir,

its velocity slows, reducing the diffusion of oxygen from the

air into the surface water. In turn, biological and chemical

oxygen demands may deplete oxygen, especially near the

bottom. This phenomenon has been well studied, and

detailed models that quantify this effect have been developed. Anoxic conditions generally cause secondary problems in drinking water, usually taste, smell, color, and

increased concentrations of iron, manganese, and sulfide.

These problems usually do not pose health risks, but may

increase water treatment costs. Under some conditions,

impoundments can cause toxic algae blooms which can pose

health risks.

Case Study: Toxic Algae Bloom at Hebgen Lake, MT

The operation of dams can affect the likelihood of bluegreen algae blooms, which sometimes produce toxins that

have been reported to be fatal to livestock, wildlife, and

pets, and pose risk to human health. For example, Hebgen

Lake, Gallatin National Forest, MT, experienced a toxic

algae bloom in June 1977 (Juday and others 1981). A family

camping at the Forest Service campground on the Grayling

Arm of Hebgen Lake (actually a medium-size reservoir) was

hysterical after their pet dog went into convulsions after

drinking some of the lake water. Their dog died a few

minutes later. When Forest Service personnel and a Gallatin

County sanitarian arrived at the campground, they were

besiged by people frightened by what they had witnessed

that day. Several more pet dogs had died, and everyone

could see the bodies of dead cattle lying near the lakeshore

beyond the campgound fence. A green scum was on the

surface of the water that was different from the algae seen

in previous years. This coating resembled thick, green pea

soup, was odorless, and went at least 50 feet (15 m)

offshore. Water samples were taken, including the green

algae, and packed in ice. The sanitarian posted his Area

Closed signs at the campground and it was closed down that

day.

The next day the samples were taken to the State Water

Quality Bureau scientists in Helena. After they heard what

had been found, they agreed to go to Hebgen Lake with

Forest Service personnel. They phoned some toxic algae

experts and reported this episode. These experts arrived a

few days later and began intensive studies of the algal

bloom. They identified the culprit as Anabaena flos-aquae, a

blue-green alga that sometimes produces a very potent toxin

(anatoxin-a), which is released into the water. No human

deaths have been attributed to anatoxin-a poisoning, but

over the past 100 years, the number of domestic and wild

animal deaths from A. flos-aquae poisoning has sometimes

numbered in the thousands. With the Fourth of July holiday

approaching, a meeting was held to decide what protective

measures should be implemented to prevent any more loss

of pets, or cattle, or risks to people. The decision was to

close the lake to recreational boating, and to keep the

shoreline and campground on the Grayling Arm of the lake

closed until the toxicity of the water had ended. Daily

sampling of the Grayling Arm algae and water continued.

The bloom gradually declined during July and was nontoxic

by July 30th.

Possible explanations for the bloom include starting to fill

the reservoir in February instead of the normal late April

because of low winter snowfall and expected low snowmelt

runoff that year, with subsequent early warming of the

water. The 21-feet (6.4-m) drawdown of this reservoir may

have allowed for bottom sediments of the Grayling Arm to

be extracted for nutrients. The upper watershed lies inside

Yellowstone National Park where it drains highly mineralized volcanic materials and geysers that produce a naturally

high concentration of nutrients. As a result, phosphate

content is relatively high. The reservoir is nitrogen limited.

Juday and others (1981) classified the main part of Hebgen

Lake reservoir as mesotrophic and the Grayling Arm as

eutrophic. They also found the A. flos-aquae algae disappeared about 1 km out in the main part of the lake. Apparently the water chemistry outside Grayling Arm was

inhospitable to the Anabaena.

If dam owners begin to fill their reservoirs earlier than

normal in the spring to capture snowmelt runoff in drought

years, the water has extra time to warm up. With enough

nitrogen and phosphorus in the warm water from natural and

manmade sources, conditions favor algae blooms. In many

States, including Montana, both Dakotas, Indiana, Iowa,

Minnesota, Missouri, and Wisconsin, toxic blooms of bluegreen algae have been reported, even in forested and largely

Hydromodifications—Dams, Diversions, Return Flows, and Other Alterations of Natural Water Flows
57

pristine watersheds (Carmichael 1981, Fawks and others

1994, Horpestad and others 1978). Whether a given bloom

will turn toxic is still unknown. Accidental ingestion by

people engaging in water sports is a risk to human health.

Although no deaths have been reported, prudence calls for

prohibiting all water contact sports and closure of public

drinking water intakes when toxic blue-green algae blooms

are suspected. Improved methods of detection of toxic bluegreen algae blooms have resulted in more reports on their

occurrence.

Trihalomethane

Trihalomethanes are compounds that form when chlorine or

bromine, added to drinking water for disinfection, reacts

with certain naturally occurring organic molecules

(trihalomethanes precursors). Trihalomethanes may cause

cancer and genetic mutations in humans. Researchers

(Arruda and Fromm 1989, Martin and others 1993) report

that reservoir and lake organic sediments can contain and

release thihalomethane precursors. In one study in Ohio, all

sediment samples had significantly more trihalomethane

precursor releases than controls. Anaerobic conditions and

deep water sediments had much fewer trihalomethane

precursors than aerobic sediments from shallower zones.

Karimi and Singer (1991) and Wardlaw and others (1991)

reviewed the role of algae as trihalomethane precursors.

They found that a variety of natural organic compounds,

especially humic and fulvic acids derived from soils and

decomposition of plant material, are the trihalomethane

precursors. No discernible trends in the ability of particular

algae species to generate trihalomethanes can be drawn from

published data. Trihalomethane concentrations arising from

a natural algal bloom, however, could theoretically exceed

maximum allowable concentrations for drinking water.

Understanding trihalomethane precursor sources is important because limiting them may lower risks to human health

and lower water treatment costs. Management of a reservoir

to limit algal growth may reduce water treatment costs and

improve water quality in the reservoir (Kortmann 1989).

Sediments Deposited in Reservoirs

Sediment deposited in reservoirs can also pose public health

problems if it contains heavy metals, radioactive elements,

or pesticides and other synthetic organic compounds. Many

of these chemically bond to the sediment particles under the

right chemical conditions. The risk to human health often

remains low as long as the sediment remains undisturbed at

the bottom of a lake or reservoir. The accidental failure or

deliberate removal of a dam may pose a human health

problem by destabilizing accumulated sediment, but

literature is lacking on this topic. Modifying streamflows

has the potential to mobilize and later deposit sediment that

may then reduce the quality of drinking water. See chapter 2

for more information on this topic. Further research needs to

be conducted on remobilization of toxic sediments.

Splash dams and log flumes were constructed on many

rivers in New England, the Lake States, and the West. The

dams were earthern structures <20 feet (7 m) high with the

main spillway constructed of wooden boards. They typically

held from a few hundred to 1,000 acre-feet (up to 1.25

million m3) of water. When the boards were removed, an

artificial flood was created downstream, sweeping logs

down the channel. Such dams are no longer constructed, and

their residual effects upon drinking water quality today are

likely to be minor.

Controlled removal of sediment by dredging from channels,

lakes, or reservoirs can degrade domestic water supplies.

These sediments pose special problems if they contain toxic

substances or if they are massively released.

Water Diversion Structures and Water

Import/Export Between Watersheds

Water is frequently removed from a river by means of a

diversionary dam or headgate along one side of the channel.

The water then enters a ditch, aqueduct, or pipeline to be

carried to the place of use, often miles away. The removal of

this water results in changes in the remaining river water.

Concentrations of pollutants increase, water temperatures

rise, and biological activity of aquatic organisms increases.

The acidity of the water often rises as well, changing the

solubility of metals and rates of chemical reactions in the

water column. Suspended sediment transport declines as

flow declines, causing increased deposition of fines on the

beds of rivers (Heede 1980).

The effects of removing water from rivers upon drinking

water quality at intakes located below points of diversion

can usually be overcome at the water treatment plant—as

long as there is enough water left to be treated. There is no

scientific literature on this subject. The same is true for

water added to stream channels by diversions from other

watersheds or aquifers. Differences in chemical, physical, or

microbiological quality of such waters may create complications when they are mixed together.

Chapter 5
58

Water Well Effects on Drinking

Water Quality

High pumping rates from water wells can decrease flows of

nearby streams used for drinking water, sometimes for

months or longer. Decreases in streamflow usually degrade

drinking water quality by changing acidity, dissolved

oxygen, and water temperature. Rates of pumping that

exceed the recharge rate of the aquifer draw down the water

table, altering the yield and water quality at other wells

tapping the same aquifer.

Wells in floodplains can become contaminated during high

streamflows if they are not properly protected ahead of time.

Singer and others (1982) found that bacterial counts, nitrate

nitrogen, turbidity, conductivity, sulfate, chloride, phosphate, total organic carbon, and several ratios of these

variables were the best indicators of surface water contamination of aquifers in a karst area of southestern Minnesota.

Improper sealing or grouting of the annular space of the well

itself can result in cross contamination, aquifer damage, loss

of well performance, and damage to the well (Ashley 1987).

The most commonly used sealing materials in wells, cement

and bentonite clay, have properties that can cause them to

fail if unsuitable drilling and well construction methods are

employed in some hydrogeologic environments. There is a

large body of literature on well construction and maintenance. The reader should obtain expert assistance if it

appears that local water wells could be responsible for

pollutants in forest or grassland watersheds. An Internet site

to go to for information on wellhead protection is EPA’s

Office of Ground Water and Drinking Water, located at

http://www.epa.gov/OGWDW/whpnp.html.

Sewage Effluent and Sludge/Biosolids

Applications to Forest and Rangeland

Return flows of sewage effluent or sludge and biosolids or

both are sometimes applied to the land surface rather than

returned to water bodies. Research on effluent and sludge or

biosolids applications was conducted in the Pacific Northwest by Machno (1989), in New England by Koterba and

others (1979), and in the Lake States and Southeastern

United States by other researchers. Materials were applied

under hardwood forests. Koterba and others (1979) found

little change in soil water and stream chemistry after light

application [11 tons per acre or 25 metric tonnes (Mg) per

hectare] of limed and dewatered sludge on sandy loam soils

in a northern hardwood forest in central New England. They

measured short-lived and relatively small increases in

calcium, magnesium, sodium, chloride, and sulfate after

56 tons per acre (125 Mg per hectare) were applied. There

were no changes in infiltration capacity of the soil and no

visual evidence of overland transport of the sludge.

Brockway (1988), Brockway and Urie (1983), and Sorber

and Moore (1986) studied effects of applying municipal or

papermill sludge and wastewater to forests by monitoring

the movement of nitrogen and other constituents in the

leachate and ground water. Results showed nitrate nitrogen

concentrations exceeded 10 parts per million in ground

water under aspen (Alnus spp.) plots treated once with 7 or

more tons per acre (16 Mg per hectare) of undigested

papermill sludge, and under pine (Pinus spp.) plantations

receiving 8.5 tons per acre (19 Mg per hectare) per year of

anaerobically digested municipal sludge in a single application. Brockway and Urie (1983) estimated that anaerobically

digested municipal sludge could be safely applied to a red

pine (P. resinosa Ait.) and white pine (P. albicaulis Engelm.)

plantation at 7.25 dry tons per acre {880 pounds total

nitrogen per acre [986 kilograms (kg) per hectare]} per year

or less, and to aspen stands at rates up to 8.4 dry tons per

acre [1,015 pounds total nitrogen per acre (1138 kg per

hectare)] per year. Although long-term additions of nitrogen

to soil could lead to nitrogen saturation (see chapter 3), this

effect has not been studied for sewage sludge applications.

Spray applications of treated municipal wastewater on

forests in Michigan have been studied by Urie and others

(1990) and Brockway (1988). Overall, it appears that nitrate

contamination of ground water can be avoided at appropriate application rates on most acidic forest soils.

Edmonds (1976) studied the survival rate over 3 years of

coliform bacteria in sewage sludge applied to a forest

clearcut on gravelly glacial outwash soils. Results indicated

that few viable fecal coliforms penetrated deeper than 2

inches (15 centimeters) into the soil and that practically

none moved into the ground water. The soil was effective as

a biological filter for hazardous pathogens, but coliforms

can remain viable for years in the surface soil. He concluded there was little danger of ground water contamination from vertical bacterial movement.

Harris-Pierce and others (1995) applied sewage sludge on a

semiarid grassland in Colorado. They found that increasing

rates of single applications from 0 to 9.7 to 18 tons per acre

(22 to 40 Mg per hectare) increased concentrations of

sediment, organic nitrogen, ammonia nitrogen, potassium,

boron, phophorous, copper, nickel, and molybdenum in

surface runoff from a single sprinkler rainfall event on the

plots. All constituents remained below EPA’s drinking water

standards. However, Burkhardt and others (1993) argued for

a careful approach to sludge applications on rangeland

Hydromodifications—Dams, Diversions, Return Flows, and Other Alterations of Natural Water Flows
59

without irrigation because of the limited opportunity for

nutrient uptake and sludge assimilation by the native

vegetation. They saw risks of the nutrients and metals

moving off-site when rainfall events do occur.

Sagik and others (1979) evaluated microbial survival and

movement in soils subjected to sludge applications and

concluded that both bacteria and viruses can survive and

move through the soil profile for up to 2 years; prudence

says that nondisinfected sludge should not be applied to

soils used to grow crops or feed for dairy cows or livestock

for human consumption. See EPA’s Web site at http://

www.epa.gov/owm/bio.htm for additional information about

biosolids recycling.

Wetland Drainage

Effects of wetland drainage on drinking water quality have

been studied. Results show some small increases in nitrogen

leaching and coliform movement with the leachate, but that

the drainage water is easily handled by the water treatment

plant.4

Reclaimed Water and Return Flows

After use, water withdrawn from rivers or aquifers is often

returned to these sources. Quantity and quality of the

returned water may be changed, depending upon the type of

use and type of treatment it receives prior to return. There is

a large body of literature and regulations about sewage

treatment because it is a point source of pollution under the

Clean Water Act. Reuse of water effluent from sewage

treatment plants is growing in the United States and has

passed the 1-billion-gallon-per-day (4-billion-liter-per-day)

mark for both nonpotable (water not intended for human

consumption) and potable (drinkable) uses. Water reuse for

nonpotable applications, such as irrigation, lawn watering,

car washing, and toilet flushing is widely accepted where

water supplies are scarce, as in Arizona, California, Florida,

and Texas. The EPA and the National Academy of Sciences

have recommended limits for many physical parameters and

chemical constituents of nonpotable water. The health risks

from disease-causing microorganisms are not as well

known; hence, there is no direct potable reuse in the United

States (Crook 1997).

Of course, indirect potable reuse occurs when effluents are

treated and returned to rivers that are water sources downstream. Required treatments may include: (1) chemical

clarification and two-stage recarbonation with intermediate

settling, multimedia filtration; (2) activated carbon adsorption; (3) ion exchange for nitrogen removal; and

(4) breakpoint chlorination. Indirect potable reuse can also

occur when effluent is used for ground water recharge by

means of injection wells. Some States prohibit that practice

if potable aquifers would be contaminated. Other States

have set stringent water-quality standards and require high

levels of effluent treatment before it is returned to the

aquifer (Crook 1997). Crook also lists a number of references on water reuse that would be very helpful to managers

of land influenced by water reuse or officials responsible for

completing source water assessments.

While irrigation return flows are exempt from the National

Pollutant Discharge Elimination System permitting process,

they can carry potentially harmful concentrations of

pesticides, heavy metals, or other toxic substances acquired

from atmospheric deposition, soils, and plants. Crop

irrigation is beyond the scope of this report; there is a large

amount of literature on this subject by EPA and various

universities.

Reliability and Limitations of Findings

Scientific literature on the direct effects of dams, water

diversion and conveyance structures, water wells, and other

related engineered structures upon drinking water quality is

very limited. Far more is known about their effects on

physical habitats of aquatic life forms. Most of the studies

mentioned did not describe how the water facility was

operated or whether the manner of operation could have, or

did, influence the results. Facility operational details should

be better evaluated in future research studies.

The indirect effects of dams, water diversions and conveyance structures, water wells, and applications of sewage

sludge should apply in all forest and rangeland watersheds

in the United States. The magnitude and timing of the

indirect effects will vary by region and perhaps by elevation

because of variations in temperature and precipitation. None

of the studies reported were national or even regional in

scope, and only a few were carried out for a decade or more,

so long-term trends have been ignored or are not known.

4

 Personal communication. 1999. James D. Gregory, Professor of

Watershed Hydrology, Department of Forestry, North Carolina State

University, Raleigh, NC 27695.

Chapter 5
60

Research Needs

1. The direct effects of dams and their operation, mobilization of sediments when dams are removed, water diversion and conveyance structures, and water wells upon

human drinking water quality need to be studied.

2. Research is needed to determine why some blue-green

algae blooms turn toxic and how to predict the toxicity

levels.

Key Points

1. Managers who experience blue-green algae blooms in

their reservoirs need to recognize that such blooms

sometimes become toxic without prior warning or

previous history. These toxins are invisible when released

by the algae into the water, and are extremely deadly to

all mammals if ingested. Most of the other risks to human

health from water storage and control structures are

known and can be assessed in local watersheds by

professionals in hydrology and health sciences.

2. Risks from applying sewage sludge on forest and

rangelands are manageable as long as disease-causing

organisms have been killed at the sewage treatment plant

before the sludge is applied to the soil.

3. Improper construction or inadequate well head protection

of water wells can be a cause of ground water contamination. People doing source water assessments in forest and

rangeland watersheds should carefully examine wells in

the vulnerability assessment.

Literature Cited

Arruda, J.A.; Fromm, C.H. 1989. Relationships among trihalomethane

formation potential, organic carbon and lake enrichment. Environmental

Pollution ENPOEK. 61(3): 199–209.

Ashley, J.W. 1987. Use and placement of grouting and sealing materials.

In: Proceedings of the 4th annual eastern regional ground water

conference; 1987 July 14–16; Burlington, VT. Dublin, OH: National

Water Well Association: 735–748.

Brockway, D.G. 1988. Forest land application of municipal sludge.

Biocycle. 29(9): 62–68.

Brockway, D.G.; Urie, D.H. 1983. Determining sludge fertilization rates

for forests from nitrate-N in leachate and ground water. Journal of

Environmental Quality. 12(4 ): 487–492.

Burkhardt, J.W.; Miller, W.W.; Azad, M. 1993. Biosolids application to

rangelands. Water Environment and Technology. 5(5): 68–71.

Carmichael, W.W. 1981. Freshwater blue-green algae (cyanobacteria)

toxins—a review. In: Carmichael, W.W., ed. The water environment:

algal toxins and health.: Proceedings of an international conference on

toxic algae; 1980 June 29–July 2; Dayton, OH. New York: Plenum

Press. 491 p.

Crook, J. 1997. Water reuse comes of age. In: Proceedings of water

resources management: preparing for the 21st century; 1997 August

10–13; Seattle. Denver: American Water Works Association: 487–501.

Edmonds, R.L. 1976. Survival of coliform bacteria in sewage sludge

applied to a forest clearcut and potential movement into ground water.

Applied and Environmental Microbiology. 32(4): 537–546.

Fawks, D.J.; Kotak, B.G.; Jones, J.R.; Prepas, E.E. 1994. Occurrence of

hepatotoxic blue-green algae in Missouri reservoirs. Lake and Reservoir

Management. 9(2): 72.

Getches, D.H.; MacDonnell, L.J.; Rice, T.A. 1991. Controlling water

use—the unfinished business of water quality protection. Boulder, CO:

University of Colorado School of Law, Natural Resources Law Center.

12 p.

Harris-Pierce, R.L.; Redente, E.F.; Barbarick, K.A. 1995. Sewage

sludge application effects on runoff water quality in a semiarid

grassland. Journal of Environmental Quality. 24(1): 112–115.

Heede, B.H. 1980. Stream dynamics: an overview for managers. Gen.

Tech. Rep. RM–72. Fort Collins, CO: U.S. Department of Agriculture,

Forest Service, Rocky Mountain Research Station. 26 p.

Horpestad, Abe; Bahls, L.L.; Glasser, S.P. 1978. Green creatures of the

deep—the Hebgen Lake algae bloom of 1977. Montana Science. 1(1):

33 p.

Juday, R.E.; Keller, E.J.; Horpestad, A. [and others]. 1981. A toxic

bloom of anabena flos-aquae in Hebgen Lake Reservoir, Montana, in

1977. In: Carmichael, W.W., ed. The water environment: algal toxins

and health. New York: Plenum Press. 491 p.

Karimi, A.A.; Singer, P.C. 1991. Trihalomethane formation in open

reservoirs. Journal of the American Water Works Association. 83(3):

84–88.

Kortmann, R.W. 1989. Raw water quality control: an overview of

reservoir management techniques. Journal of the New England Water

Works Association. 103(4): 197–220.

Koterba, M.T.; Hornbeck, J.W.; Pierce, R.S. 1979. Effects of sludge

applications on soil-water solutions and vegetation in a northern

hardwood stand. Journal of Environmental Quality. 8(1): 72–78.

Machno, P.S. 1989. Seattle’s success with silvigrow broadens sludge

management options. In: Proceedings of the national conference; 1989

May 27–29; Boston. Sludge Manage. Ser. 17. Silver Spring, MD:

Hazardous Materials Control Research Institute: 131–134.

Martin, A.B.; Cooke, G.D.; Carlson, R.E. 1993. Lake sediments as

potential sources of trihalomethane precursors. Water Resources

Research. 27(12): 1725–1729.

Reetz, G.R.; Berkley, J.; Campbell, D. [and others]. 1998. Water quality

in the West. Report to the western water policy review advisory

commission. Denver: U.S. Environmental Protection Agency. 183 p.

+ appendices.

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61

Sagik, B.P.; Moore, B.E.; Sorber, C.A. 1979. Public health aspects related

to the land application of municipal sewage effluents and sludges. In:

Sopper, W.E.; Kerr, S.N., eds. Utilization of municipal sewage effluent

and sludge on forest and disturbed land. University Park, PA: The

Pennsylvania State University Press: 241–253.

Singer, R.D.; Osterholm, M.T.; Straub, C.P. 1982. Ground water quality

in southeastern Minnesota. Water Resour. Res. Cent. Bull. 109. St. Paul,

MN: University of Minnesota. 45 p. Available from the National

Technical Information Service, 5285 Port Royal Road, Springfield, VA

22161; PB83–15437.

Sorber, C.A.; Moore, B.E. 1986. Microbiological aspects of forest

application of wastewater and sludge. In: Cole, D.W.; Henry, C.L.;

Nutter, W.L., eds. The forest alternative for treatment and utilization of

municipal and industrial wastes. Seattle: University of Washington

Press: 73–84.

Urie, D.H.; Hart, J.B.; Burton, A.J. 1990. Nitrification in sludge—

amended Michigan forest soils. Journal of Environmental Quality. 19(3):

609–616.

U.S. EPA. 1995. National water quality inventory: 1994 report to Congress.

EPA–841–R–95–005. Washington, DC: U.S. Environmental Protection

Agency. 497 p.

Wardlaw, V.E.; Perry, R.; Graham, N.J.D. 1991. The role of algae as

trihalomethane precursors: a review. Aqua. 40(6): 335–345.

Chapter 5
62

Chapter 6

Urbanization

Wayne C. Zipperer, Karen Solari, and Beverly A. Young1

Introduction

This chapter specifically examines drinking water issues

related to urbanization. The discussion is limited principally

to land that is developed or being developed within and

adjacent to public land. Urbanization issues include current

and past land uses and Forest Service facilities. Forest

Service buildings and administrative sites are included

because they are similar to other developed sites.

The effects of urbanization on drinking water quality

encompass many topics with extensive published literature.

Because of limited space, selected topics of current and past

land use are examined including wastewater treatment,

urban storm runoff, underground storage tanks, abandoned

wells, and landfills. However, land managers need to realize

that these selected topics do not include all effects caused by

urbanization, land fills, and abandoned mines on drinking

water and serve only to illustrate potential effects.

Vitousek (1994) has identified land cover changes as a

primary effect of humans on natural systems. With the

projected global increase in urbanization, land cover

conversions for urban use will only increase. In this chapter,

we examine the potential impacts on drinking water of

current and past use of land in and adjacent to public land.

Nationally, development and growth rates are not available

for such land. To estimate these rates, we have used inholding data from the Forest Service.

Inside the boundaries of publicly owned land are parcels not

administered by the Agency. They are called in-holdings. Inholdings are managed or owned by other Federal, State,

local, and tribal government agencies, and by private

landowners. Of particular interest is the private land because

of its propensity for development. The occurrence of inholdings varies by Forest Service region (table 6.1). The

States that comprise each region are listed in table 6.1.

Region 4 has the least area of in-holdings; only 6.9 percent

of the land inside national forest boundaries. By comparison, Regions 8 and 9 had 48.6- and 45.6-percent in-holdings

within national forest boundaries, respectively.

Unfortunately, no data are available on how rapidly these inholdings are being developed. To estimate this rate, we used

the growth rates of counties that intersect with or are

adjacent to a national forest. Population growth was

calculated for 1980–90 and 1990–96 using census data (U.S.

Census Bureau 1997). Between 1980 and 1990, the population of these counties grew by 18.5 percent, while the

Nation’s population grew by 9.8 percent (table 6.1). For

1990 to 1996, the population in these counties grew by 10.1

percent, while the Nation’s population grew by 6.4 percent.

In 1996, these counties contained 22.4 percent of the

Nation’s population (U.S. Census Bureau 1997).

Population change in these counties varied by region and

time period. Between 1980 and 1990, Region 5 experienced

the greatest percent increase (28.9 percent), while Region 1

experienced a decrease of 3.5 percent. Between 1990 and

1996, Region 4 and 6 showed the largest increases of 14.9

and 14.3 percent, respectively, and Region 9 showed the

least growth of 5.2 percent. Overall, Regions 2, 3, 4, 5, and

6 had growth rates greater than the national average for the

period between 1990 and 1996. The effects of this development on drinking water quality depend on the location

within the watershed, the concentration of development, and

existing conditions. Unfortunately, data are unavailable to

examine those variables.

Issues and Risks

During the past 20 years, private tracts in and adjacent to

public land have been developed rapidly for residential,

commercial, and recreational use. This development poses a

significant threat to drinking water quality through surface

and ground water contamination. Development occurs near

the headwaters of streams where water quality is generally

the highest and is easily degraded because of stream size.

For ground water, about 95 percent of rural communities use

1

 Research Forester, USDA Forest Service, Northeastern Forest Research

Station, Syracuse, NY; Environmental Engineer, USDA Forest Service,

Washington, DC; Environmental Engineer, USDA Forest Service, Northern

Region, Missoula, MT, respectively.
63

ground water as the principal source of drinking water.

Sources of pollution result from wastewater treatment,

nonpoint-source pollution, underground storage tanks, solid

waste storage, and hazardous material storage. The extent of

ground water contamination depends on depth of ground

water. Shallow ground water sources <100 feet [30 meters

(m)] below land surface may be more readily and significantly contaminated than deeper ground water sources (U.S.

Geological Survey 1999).

In 1995, the U.S. Environmental Protection Agency (EPA)

(U.S. EPA 1998d) summarized water-quality information

submitted by States, tribes, and other jurisdictions. For

rivers, streams, lakes, ponds, and reservoirs, municipal point

and nonpoint sources from residential and commercial

sources were identified as significant contributors of

pollution to rivers, streams, lakes, ponds, and reservoirs. For

ground water, principal sources included leachate from

leaking underground storage tanks, septic tanks, and

landfills. Of specific importance is the effect of urbanization

on the quality of surface water and ground water in rural

areas (table 6.2).

The report identified urbanization as a major factor in

contaminating surface and ground water, and modifying

hydrologic processes. Urbanization replaces natural vegetation cover with impervious surfaces, decreasing natural

infiltration of water, increasing peak flows, and decreasing

ground water recharge (Weiss 1995). Increased peak flows

can negatively affect drinking water quality by causing bank

destabilization and streambed scouring, which increase

turbidity and sedimentation (Phillips and Lewis 1995).

Reduced ground water recharge decreases baseflow in

streams and increases pollutant concentrations. Decreased

baseflow impairs aquatic habitat and riparian wetlands and

increases the stream’s sensitivity to pollution and sedimentation (Weiss 1995).

Table 6.1—Reported acres inside national forest boundaries and percent population changes in counties

containing or adjacent to national forests by administrative Forest Service Region in the conterminous

United States

Population Population

change change

Regiona Gross NF In-holding Other 1980–90 1990–96

- - - - - - - - - - - - - - - - - -Acres - - - - - - - - - - - - - - - - - - - - - - - - - - - - Percent - - - - - - - - - -

1 28,180,534 25,375,333 2,805,201 10.0 -3.5 5.7

2 24,477,648 22,098,044 2,379,604 9.7 3.6 10.6

3 22,381,905 20,702,312 1,679,593 7.5 17.2 10.9

4 34,257,094 31,903,934 2,353,160 6.9 14.8 14.9

5 23,739,894 20,022,650 3,717,244 15.7 28.9 10.3

6 27,357,569 24,629,048 2,728,521 10.0 6.8 14.3

8 25,034,868 12,874,851 12,160,017 48.6 6.1 6.8

9 21,934,418 11,942,218 9,992,200 45.6 1.8 5.2

Total 207,363,930 169,548,390 37,815,540 18.2c 9.8d 6.4d

a Region 1: Montana, northern Idaho, North Dakota, and northwestern South Dakota; Region 2: Colorado, Kansas, Nebraska, and

southeastern South Dakota; Region 3: Arizona and New Mexico; Region 4: southern Idaho, Nevada, Utah, and western Wyoming;

Region 5: California, Hawaii, Guam, and Trust Territories of the Pacific Islands; Region 6: Oregon and Washington; Region 8: Alabama,

Arkansas, Florida, Georgia, Kentucky, Louisiana, Mississippi, North Carolina, Oklahoma, Puerto Rico, South Carolina, Tennessee,

Texas, and Virginia; and Region 9: Connecticut, Delaware, Illinois, Indiana, Iowa, Maine, Maryland, Massachusetts, Michigan, Minnesota,

Missouri, New Hampshire, New Jersey, New York, Ohio, Pennsylvania, Rhode Island, Vermont, West Virginia, Virgin Islands, and

Wisconsin.

b In-holding = parcels of land within the boundaries of publicly owned lands that are not administered by the public agency;

NF = national forest; Other = percent of in-holdings within the gross acreage inside national forest boundaries.

c Percent of in-holdings within national forest boundaries for the conterminous United States (calculated as in-holdings/gross).

d Percent population change in counties containing or adjacent to national forests for the conterminous United States.

Chapter 6

Areab
64

Wastewater Treatment

Residential and commercial wastewater is treated by

decentralized and centralized systems. Decentralized

systems treat water onsite. They include individual and large

septic systems, and cluster wastewater systems. Generally,

septic systems treat and dispose of relatively small volumes

of wastewater. They are for individual dwellings or groups

of dwellings and businesses located close together. A

centralized system is a collection and treatment system

containing collection sewers and a central treatment facility

(U.S. EPA 1997a). Centralized systems are used to collect

and treat large volumes of water. Decentralized systems

affect both surface and ground water, while centralized

systems generally affect surface waters.

The 1990 census indicates that 25 million households use

onsite disposal systems for wastewater. Data on the failure

rates associated with these systems are limited and no

national estimates are available. Each State has its own

definition of failure, but estimates of failure rates range from

18 to over 70 percent (U.S. EPA 1997a). Twenty-seven

States have cited onsite disposal systems as a potential

source of ground water contamination (U.S. EPA 1998d).

Contaminants from onsite disposal can be classed as

inorganic (sodium, chlorides, potassium, calcium, magnesium, sulfates, and ammonium), microorganisms (bacteria

and viruses), and chemical organics originating in household

products (Phillips and Lewis 1995, U.S. EPA 1997a).

Effluent from septic systems usually contains high concentrations of ammonium and organic nitrogen.

Water supplies are vulnerable to pathogenic bacteria and

viruses from onsite disposal systems. Reported outbreaks of

waterborne disease in the United States are uncommon

(table 6.3), but 404,000 people fell ill to a Cryptosporidum

spp. outbreak in Milwaukee, WI, in 1993. To some extent,

low occurrence may be attributed to individuals being

unaware that their illness was a waterborne disease or to the

number of illnesses being so small that they go unreported

by local health departments. Ground water sources have a

higher incidence of waterborne outbreaks than surface water

because ground water often is not filtered or disinfected

before it is used for drinking (table 6.4). Disease-causing

microorganisms isolated from domestic sewage include

Salmonella, Shigella, pseudomonas, fecal coliform, and

protozoa (Giardia lamblia) (U.S. EPA 1997a). Other

microorganisms found in contaminated drinking water

include Cryptosporidum, Microsporidum, Cyclosporidium,

Helicobacter pylori, hepatitis E, and the enteric viruses

hepatitis A and Norwalk virus (U.S. EPA 1997b). See

chapter 2 for a more thorough discussion of waterborne

pathogens.

Table 6.2—Estimated use from freshwater surface

and ground water sources in the United States,

1980–95

Source 1980 1985 1990 1995

 - - - - - - - - Cubic kilometers - - - - - - -

Ground 120 101 110 105

Surface 400 366 358 364

Total 520 467 468 469

Source: Adapted from Gleick 1999.

Urbanization

Table 6.4—Comparison of outbreak percentages by

drinking water source from pathogenic contamination

for the period 1971–96a

Water source Total outbreaks Cases of illnesses

No. % No. %

Ground 371 58 84,408 52

Surface 215 33 66,721 41

Other 56 9 10,625 7

a Excludes outbreak in Milwaukee, WI, 1993.

Source: Adapted from Craun and Calderon 1996, U.S. EPA 1999.

Table 6.3—Waterborne disease outbreaks in the United

States by water supply system, 1990–94

Source 1990 1991 1992 1993 1994

Municipal 5 2 9 9 5

Semi-public 7 13 14 4 5

Individual 2 0 4 5 2

Total outbreaks 14 15 27 18 12

Total cases 1,758 12,9604,724 404,190a 1,176

a Includes Milwaukee, WI.

Source: Adapted from Gleick 1999.
65

The fate and transport of parasites, bacteria and viruses from

sewage effluent depend on the characteristics of the subsurface environment (U.S. EPA 1997a). Pore size and chemical

charges of the soil matrix are important in removing bacteria

and viruses. Bacteria have been reported to travel distances

of up to 300 feet (100 m) in sand aquifers, 2,500 feet (800

m) in gravel aquifers, and over 3,000 feet (1000 m) in

limestone rock (Kaplan 1991). Certain viruses, because of

their size and long survival times, can travel distances up to

1 mile [1.6 kilometers (km)] in areas with karst geology

(Yates and Yates 1989).

Organic chemicals in onsite disposal systems are a less

commonly reported problem because they often are below

levels considered hazardous to human health (U.S. EPA

1997a). These chemicals can significantly affect aquatic

systems, however. Organic chemicals commonly found in

septic systems originate from household products, paints

and varnishes, shampoos, cosmetics, and polishes.

Septic systems fail for two reasons: poor design or poor

maintenance (Kelley and Phillips 1995). Design includes

construction, soils and hydrological characteristics of the

site, and drainfield layout (Kelley and Phillips 1995). If

drainage is too slow, there will be upward seepage and

ponding, which are likely to contaminate surface water. If

drainage is too fast, downward percolation occurs without

sufficient biological treatment; contamination of ground

water is likely to result.

Even with properly installed systems, maintenance is

absolutely necessary. Unfortunately, the typical owner of an

onsite disposal system is unaware of the need for proper

maintenance (Kelley and Phillips 1995). Maintenance

includes periodic testing of drainfields and emptying of

septic tanks. Frequency of maintenance depends on soil

conditions, type of septic system, and weather patterns.

Class V injection wells is another type of onsite disposal

unit. An injection well can include any manmade hole in the

ground for injection of wastewater (U.S. EPA 1998a). They

are used by dry cleaners, laundromats, paint dealers,

hardware stores, funeral homes, and other industrial and

commercial facilities for materials other than domestic and

sanitary wastes. Motor vehicle waste disposal wells,

industrial waste disposal wells, and large-capacity cesspools

have high risk for ground water contamination. Field studies

have shown that ground water sources can be degraded

significantly by organic and inorganic contaminants from

dry wells in automotive shops (Ogden and others 1991).

See the section on abandoned wells in this chapter.

Approximately 10 percent of the wastewater produced

in the United States originates from communities of

<10,000 people. With the passage of the Clean Water Act in

1972, many such communities elected to use Federal funds

to install centralized systems for wastewater treatment. In

small communities, contractors frequently installed the most

economical and not necessarily the most effective systems.

Currently, many of these systems are obsolete and need

replacing because they have operated beyond their 20–year

life span. Small communities also face an economic factor

of scale. Costs of maintenance and staffing must be divided

among fewer people, resulting in higher costs per person.

Consequently, small communities have nearly twice the

number of violations than larger communities (>10,000

individuals). Violations include leaking sewage systems

(cracked and broken sewer lines), illegal connections of

sewer and storm drainage lines, and inadequate treatment.

Violations often affect local water quality and potentially

affect drinking water quality for downstream communities.

Since 1970, new technologies have been developed to treat

water more effectively and cheaply. However, many small

communities have not adopted these systems because of a

lack of knowledge, public distrust of new technologies, and

legislative and regulatory constraints (U.S. EPA 1994).

Additional discussion of centralized wastewater treatment

can be found in chapter 5.

Urban Runoff

Findings

Urban land generates nonpoint-source pollution. People

apply various chemicals around their homes, businesses, and

adjacent land. These chemicals are carried by surface runoff

to receiving waters. As land is developed and impervious

surface area increases, the amount of urban runoff increases.

Consequently, land development increases the amount of

nonpoint-source pollutants discharged into surface water

(Phillips and Lewis 1995). The Nationwide Urban Runoff

Program (U.S. EPA 1983) reported that 10 times as much

suspended solid material was being discharged from storm

sewers serving residential and commercial areas as was

discharged from sewage treatment plants providing secondary treatment (Weiss 1995). Major pollutants associated

with residential and commercial runoff include floatables,

sediments, suspended solids, oxygen-demanding materials,

nutrients, organics, biocides (herbicides, fungicides,

pesticides), polycyclic aromatic hydrocarbons, and petroleum hydrocarbons (U.S. EPA 1997a, Weiss 1995).

Because residential and commercial construction creates site

disturbances, it is highlighted here. Sediment loading from

site preparation, and construction and maintenance of

buildings and roads can exceed the capacity of streams to

Chapter 6
66

transport it (Yoder 1995). Sediment loads from inadequately

controlled construction sites typically are 10 to 20 times per

unit of land area those from agricultural land and 1,000 to

2,000 times those from forest (Weiss 1995). In a relatively

short period, urban site construction can contribute more

sediment to a stream than was deposited over the previous

several decades (Weiss 1995).

Urban runoff is highly intermittent. Short-term loading,

associated with individual storms, is high and may have a

shock effect on the receiving water (Weiss 1995). When

predicting the effect of urban runoff on water quality, it is

important to determine the duration of the effect. Effects

may be acute (short term) or chronic (long term) (Phillips

and Lewis 1995). Oxygen-demanding substances and

bacteria create acute effects; whereas nutrients, sediments,

toxic metals, and organics create chronic effects. For an

acute effect, estimates are based on the probability that

pollution concentrations will exceed acceptable drinking

water standards (Phillips and Lewis 1995). For a chronic

effect, a simple method has been developed to predict the

increase in pollution loading above current conditions (U.S.

EPA 1983). This simple method is employed by EPA and

uses information readily available to the resource manager.

Input variables include pollutant type and concentration,

precipitation, and percent impervious cover. The method,

however, is limited to areas <1 square mile (2.6 km2).

Findings from engineering research show that pollution and

sediment loading from runoff can be reduced. Practices for

mitigating storm runoff include attenuation, conveyance,

pretreatment, and treatment of runoff (U.S. EPA 1997a).

When selecting mitigation practices, it is important to

consider

• How will practices meet watershed and site objectives?

• What are the limitations of a practice to meet objectives?

• What are the drainage field, soil types, and topography?

• Are practices compatible with a region’s rainfall pattern

and annual runoff?

• Are they derived from scientific research?

• How will practices function as a system (Phillips and

Lewis 1995)?

A number of manuals and practical guides have been written

to select, design, and maintain mitigation practices to meet

local, State, and Federal mandates (Birch 1995, Phillips and

Lewis 1995). As with plans and guides for wastewater

treatment, managers need to check with State and local

agencies for specific performance ratings and regulations.

Like wastewater treatment facilities, new mitigation

practices must be maintained and existing ones upgraded to

meet expected performance standards. Adequate funds often

are lacking to maintain or enhance these facilities (U.S. EPA

1997a). Without proper maintenance, water quality degrades

as systems fail.

Reliability and Limitation of Findings

Although development of private land in or adjacent to

public land has occurred for decades, scientific studies of

the effects on water quality and drinking water sources are

lacking. Extensive research has been conducted on urban

effects on natural systems, however. These studies provide

the basis for identifying the potential impacts of development on drinking water.

When applying findings across a watershed, scale becomes

an important issue. Evaluating cumulative effects requires

examination of more than just local impacts of individual

pollution sources, such as urban runoff, wastewater treatment, and landfills. The timing and location of all activities

that contribute contaminants within the watershed and their

hydrologic connection to source water intakes must be

considered to estimate cumulative effects. Consequently,

these developments must be evaluated both independently

and collectively within the watershed. Ages of wastewater

treatment facilities and urban storm runoff structures must

be considered. For various reasons, existing infrastructures

may not meet sanitation and water-quality regulations.

Success of management plans to mitigate the effects of

wastewater treatment and urban runoff is predicated on

sound infrastructure.

The ability to address the effects of development on

drinking water quality depends on ownership. On publicly

owned land, resource managers directly determine whether

facilities comply with Federal and State regulations. On

privately owned land, resource managers can only indirectly

influence development effects on drinking water quality

through the planning process.

Planning and development of private land in and adjacent to

public land involve complex issues including the interplay

of the physical, biological, and social components of a

watershed. A number of factors need to be considered. First,

planning must include all stakeholders, including public

land managers. Second, private tracts are owned by a

diversity of individuals for various reasons. Third, new

regulations often cause resentment among landowners.

Any changes in drinking water regulations and statutes

create the need for communication and education. Fourth,

a comprehensive approach is needed to account for the

Urbanization
67

cumulative effects of individual developments in a watershed and to address the needs of individual stakeholders.

Because of the interplay, technology and management

practices are not the only solutions to drinking water issues.

A number of communities have adopted a whole watershed

approach to manage water and land planning issues (Birch

1995, Kelley and Phillips 1995, Phillips and Lewis 1995).

This approach provides a framework not only to design the

optimal mix of water-quality management strategies but also

to design land management strategies by integrating and

coordinating management priorities across stakeholders,

governments, and agencies. Livingston (1993) identifies

the big “C’s” of watershed management that must be

considered:

• Comprehensive management.

• Continuity of management over a long period.

• Cooperation among Federal, State, local, and tribal

governments; cities and counties; public and private

sectors; and all citizens.

• Communication to educate elected officials and ourselves.

• Creativity in best-management-practice technology.

• Coordination of stormwater retrofitting to reduce pollution loading.

• Consistency in implementing laws, rules, and programs.

• Commitment to solving current problems and preventing

future ones.

• Cash in funding programs and maintenance over a long

period.

Research Needs

Development of private tracts in and adjacent to public land

represents an opportunity to examine how development

alters ecosystem processes and what are the long-term

implications of these changes.

1. Long-term monitoring stations are needed not only to

monitor changes in water quality and habitat modification

but also atmospheric deposition.

2. In addition, studies are needed to determine the limitation

of management practices, wastewater treatment, and

urban runoff in extreme environments such as at high

elevations [>8,000 feet (>2,400 m)]. Research also is

needed to determine threshold levels of the corresponding

changes in processes that affect source water quality as

land use shifts to urban.

Key Points

1. Levels of drinking water protection need to increase with

increasing amounts of urban development.

2. Because of their depth, shallow ground water sources are

especially prone to contamination from septic systems.

3. Septic systems fail for two reasons: poor design and poor

maintenance.

4. Septic system designs need to consider site conditions,

such as soil characteristics (permeability, depth to

bedrock, depth to ground water table), topography

(floodplain, hillslope, ridge top), and climatic patterns

(rainfall and snowfall amounts and patterns, winter

temperatures).

5. A comprehensive approach towards development

planning must be taken. The approach needs to consider

issues ranging from the local to watershed scale.

6. Urban runoff is reduced by maintaining and enhancing

existing vegetation and by minimizing the amount of

impervious surfaces.

Underground Storage Tanks

Issues and Risks

Underground storage tanks pose a risk of ground water

contamination because nearly all tanks contain petroleum

products. The tanks are associated with service stations,

convenience stores, and organizations that have fleets of

vehicles (U.S. EPA 1998b). Current estimates indicate that

25 to 35 percent of these tanks do not comply with existing

regulations. In 1986, EPA published regulations with the

goals of preventing and cleaning up releases from underground storage tanks. These regulations (40 CFR 280)

require that underground storage tanks, which contain

hazardous substances, including fuels, be removed by

December 1998 or have spill, overfill, and corrosion

protection. The regulations also require that installation and

closure of underground storage tanks must be registered

with the State or EPA. These regulations have had a significant impact on land management agencies, which, due to the

remote locations of administrative offices, recreation sites,

and workshops, have installed underground storage tanks for

easy access to fuel. For example, in order to comply with

these regulations, the Forest Service has removed over 1,600

underground storage tanks and has initiated several projects

to cleanup contaminated soil caused by leaking tanks.

Chapter 6
68

The primary concern about underground storage tanks is

leakage, which can seep into the soil and contaminate

ground water. Since 1988, over 330,000 confirmed releases

have occurred from regulated underground storage tanks.

Gasoline is the most common contaminate of ground water.

Although not all of those releases contaminated ground

water, drinking water wells have been shut down because of

petroleum contamination (U.S. EPA 1996). In 1988, EPA

regulations established minimum standards for new tanks

and required owners to upgrade existing tanks, to replace

them, or close them by December 1998 (U.S. EPA 1996).

Findings

Recent studies have identified methyl tertiary butyl ester

(MTBE) as a potential major health hazard in drinking

water. Methyl tertiary butyl ester is added to gasoline to

increase its oxygen content and to reduce airborne emissions. Effects on drinking water include widespread impacts

from low concentrations and local impacts from high

concentrations (U.S. EPA 1998c) (see chapter 7 for more

detailed information on the effect of vehicular emissions).

Local impacts primarily result from leaking underground

storage tanks. A survey of ground water plume data from

over 700 service stations showed that 43 percent of the sites

had MTBE concentrations >1,000 micrograms (μg) per

liter. However, a survey of drinking water wells from 20

National Water Quality Assessment study units showed that

2 percent of 949 rural wells had a median concentration of

approximately 0.5 μg per liter (well below the EPA drinking

water advisory of 20 to 24 μg per liter) (U.S. EPA 1998c,

Zogorski and others 1998). A study of private wells in

Maine showed 1.1 percent of 951 wells with MTBE levels

exceeding 35 μg per liter. Maine officials estimated that

1,400 to 5,200 private wells across the State could be

contaminated at levels exceeding 35 μg per liter (U.S. EPA

1998c). The potential threat of underground storage tanks

contaminating ground water should diminish as older tanks

are upgraded and sites are cleared of contaminates.

Reliability and Limitation of Findings

Records should be available through the State or EPA

identifying where underground storage tanks are located,

where cleanup operations are ongoing, and where tanks

have been removed. The possibility also exists that underground storage tanks may be present and not registered with

the appropriate agency. During field visits, resource managers need to look for indications of former structures or

operations on the property, and they need to note the

presence of partially exposed, capped, or uncapped pipes.

These pipes may be vent pipes or fill pipes for underground

storage tanks. On properties where motor vehicles were

operated regularly, be skeptical where there is no apparent

refueling source. An underground storage tank is likely to be

present (U.S. Department of Agriculture, Forest Service

1999).

Research Needs

1. More data are needed to determine the extent of contamination of drinking water sources by MTBE and the

potential health hazard.

2. Research also is needed to develop more effective and

cost efficient cleanup methods. Cleanup of ground water

and soil contaminated by leaking underground storage

tanks can be expensive and take long periods of time.

Key Points

Underground storage tanks are a potential threat to drinking

water supplies through contamination of surface and ground

water by storage tanks that have leaked or have been

overfilled.

Abandoned Wells

Issues and Risks

Abandoned wells and wells that are no longer used may or

may not have been properly closed or plugged after their use

ceased. Abandoned wells are of concern because they can

serve as conduits for migration of contaminants into aquifers

and between aquifers.

Numerous types of abandoned wells exist on public land.

Some were drilled for mineral exploration, others for oil and

gas production, and still others for stock watering. Those

associated with administrative and recreational developments include water wells for irrigation and drinking water

and disposal wells for stormwater runoff or waste products

from vehicle shops. Septic systems may be considered

disposal wells when industrial or commercial wastes are

treated along with sanitary wastes.

Although Federal, State, and local regulations address

proper closure of abandoned wells, not all abandoned wells

have been closed or plugged properly. Many of the improperly closed wells were abandoned before regulations existed.

Other wells have been abandoned temporarily to allow for

further use if the need should arise. Certain wells, such as

automotive dry wells in vehicle shops, may still be in use

but would be banned or subject to permit under proposed

Urbanization
69

regulations for Underground Injection Control (U.S. EPA

1998e). Certain States already have banned such dry wells

and have required cleanup, per the Resource Conservation

and Recovery Act of 1976, due to contamination at such

sites.

The number of abandoned wells on public land is unknown.

For example, the Forest Service has inventories of some

categories of in-use wells but not of abandoned wells.

Knowledge of number and location of such wells is limited,

and in most instances, might be gained only by a field

survey.

Findings

Abandoned wells are commonly cited as avenues of

contamination in Federal, State, and local programs dealing

with ground water protection (Nye 1987). The EPA’s Adopt

Your Watershed campaign supports properly closing

abandoned wells. Many States, such as Iowa, Kansas, and

Nebraska, provide financial incentives for proper well

abandonment because it is considered so important for

ground water protection.

Proper abandonment of water wells is regulated at the State

or local level. Oil and gas well closure is specified in 43

CFR 3160. Motor vehicle waste disposal wells (dry wells)

are regulated in the underground injection control program

as class V wells (40 CFR 146).

Field studies have shown that ground water sources can be

degraded significantly by organic and inorganic contaminants from stormwater runoff and dry wells in automotive

shops (Ogden and others 1991). In certain geologic formations, abandoned water wells are prone to collapse, and,

when wells are drilled through multiple aquifers, contamination problems may occur (Blomquist 1984). Gass (1988)

reported that abandoned water supply wells became conduits for cross contamination between aquifers. Abandoned

oil and gas wells allowed leakage of contaminated or highly

mineralized water, leading to ground water pollution

including salinization (Gass 1988). Even plugged boreholes

may have defects in structural integrity, allowing pollutant

transport between confined aquifers (Avci 1992).

Reliability and Limitation of Findings

The issue of abandoned well closure is well defined in

Federal, State, and local regulations. The extent of the

problem on public land is unknown because wells have not

been inventoried. Proper well closure is heavily regulated at

present, but not heavily enforced. Existence of improperly

closed wells does not mean ground water contamination will

occur; only that it is has the potential to occur.

Wells on public land possess the same general characteristics as other abandoned wells. Drilling and development

methods for all types of wells have usually followed

industry standards. For all types of wells, the newer the well

the more likely that it was drilled and closed properly. On

public land, dry wells in vehicle shops may not have as

much waste or as much variety of waste in them as a

commercial facility would, but the pollution potential still

exists. Some could have greater potential for contamination

than others because of hydrogeologic formations and

duration of well use. For example, in the Allegheny and

Appalachian Mountains, where abandoned oil and gas wells

are more numerous and older, problems may be greater than

in other regions of the country.

Research Needs

1. Methods need to be developed to inventory abandoned

wells on both public and private land. Inventorying

methods need to incorporate the capabilities of remote

sensing technology and Geographical Information

Systems.

2. The inventorying process also needs to be linked to a

monitoring program.

3. Further, an abandoned-well typology needs to be

developed that integrates type of well, geological

formation, soil, typography, climate, and potential for

ground water contamination.

Key Points

1. Abandoned wells may serve as conduits for the transport

of pollutants.

2. Where there may be no records of abandoned wells on a

property, the property must be surveyed to locate wells.

3. The type of abandoned well influences the types of

pollution that may enter ground water sources.

4. Improperly sealed abandoned wells may be a source of

contamination.

Chapter 6
70

Solid Waste Landfills and Other

Past Land Uses

In 1990, citizens in the United States generated over 195

million tons [215 million metric tonnes (Mg)] of municipal

solid waste. Currently, over 6,000 regulated municipal

landfills exist (U.S. EPA 1993). However, an estimated

30,000 to 50,000 unregulated waste disposal sites are

thought to exist in the United States (Woldt and others

1998). Both regulated and unregulated sites may have

impacts on water quality and the environment. In 1976, the

Resource Conservation and Recovery Act (RCRA) addressed waste management and separated hazardous waste

management from solid waste management. Prior to RCRA,

municipal and industrial wastes were deposited at the same

landfills. The practice was to spread hazardous waste sludge

and liquids over municipal waste, using the municipal waste

to soak up the sludge (Brown and Donnelly 1988). Consequently, landfills existing prior to RCRA may contain

hazardous waste and may be the origin of organic compounds found in municipal landfill leachate. Other sources

of hazardous materials in landfills include household and

agricultural materials, incinerator ash, and sewage sludge.

The U.S. EPA (1993) defines a municipal solid waste

landfill as:

A discrete area of land or an excavation that receives

household waste, and that is not a land application

unit, surface impoundment, injection well, or waste

pile, as those terms are defined in the law. (Household waste includes any solid waste including

garbage, trash, and septic waste derived from houses,

apartments, hotels, motels, campgrounds, and picnic

grounds.) A municipal solid waste landfill unit also

may receive other types of waste such as commercial

solid waste, non-hazardous sludge, small quantities of

generator waste, and industrial solid waste.

In many rural areas, small communities are served by small

landfills that may be exempt from some regulatory requirements. The U.S. EPA (1993) defines a small landfill as one

that receives less than an average of 20 tons (22 Mg) of

waste per day, receives <25 inches (62.5 centimeters) of

rain per year, and shows no evidence of ground water

contamination. About half of the regulated landfills serve

communities with <10,000 people and are considered small

landfills. Many of these small landfills may be on or

adjacent to public land.

Issues and Risks

Municipal solid waste landfills that contaminated ground

water often were poorly designed, located in geologically

unsound areas, or accepted toxic materials without proper

safeguards (U.S. EPA 1993). Decomposing municipal solid

waste in landfills form leachates, liquids containing extremely high concentrations of organic and inorganic

pollutants. Ground water contamination is common near

landfills, but the effect may decrease with distance (Borden

and Yanoschak 1990). A study of 71 North Carolina sanitary

landfills found that 53 percent had ground water violations

for organic and inorganic pollution based on North Carolina

ground water-quality standards (Borden and Yanoschak

1990). Only a few landfills had organic contamination.

When predicting the performance of a landfill, it is important to know its age, history of material disposal, design, and

capability of handling toxic waste.

Another threat of landfills to ground water is volatile

organic compounds (VOC). Volatile organic compounds

come from biological and chemical degradation of materials

in the landfill. Recently, VOC’s have been detected in

ground water (Baker 1998) and management procedures

have been developed to minimize this threat (Rickabaugh

and Kinman 1993). Ground water contamination was linked

to methane diffusion as VOC concentrations increased.

Mitigation involves improving gas removal systems at the

landfill (Baker 1998). The extent of ground water contamination by VOC’s and subsequent health effects need to be

evaluated further.

Illegal dumping may occur on or adjacent to public lands.

This practice is usually done to avoid disposal fees or the

time and effort required for proper disposal. Dumped

materials may include nonhazardous material such as scrap

tires, yard waste, and construction waste. It also may include

hazardous waste such as asbestos, household chemicals and

paints, automotive fluids, and commercial or industrial

waste. The potential for contaminated runoff and ground

water depend on such factors as the proximity of the dump

to surface water, elevation of the ground water table, and

permeability of the soil.

Other sources of contamination on public land include

shooting ranges, formerly used defense sites, and wood

treatment sites. Shooting ranges pose the potential for lead

contaminates entering surface water and ground water.

Acidic rainfall or acidic soil can dissolve the weathered lead

compounds. In a dissolved state, lead can move through the

soil and enter surface water and ground water. Shooting

ranges in areas with acidic soils or acidic rainfall have an

increased potential for transporting contaminates offsite and

into drinking water. Bare ground on ranges may further

increase the risk of migration of lead compounds offsite.

Sites once used by the Department of Defense (DOD) for

military training and industrial facilities are on both public

Urbanization
71

and private land. The DOD estimates that over 9,000 such

sites exist. They pose a wide range of environmental

hazards, including unexploded ordinances from the training

sites and soil contamination from solvents, fuels, and other

petroleum compounds used at industrial facilities. Sites are

being cleaned up to minimize environmental effects. The

Forest Service, for example, has identified over 100

formerly used defense sites on national forests.

Field treatment of wood posts is another past land-use

activity that may have led to surface and ground water

contamination. The common practice was to dip wooden

posts into tanks that contained creosote, pentachlorophenol,

or a chromium, copper, and arsenic compound and move

them to an area for dripping and drying. The practice has

been discontinued on public land such as national forests.

However, a potential exists for surface and ground water

contamination from past wood treatment operations.

Reliability and Limitation of Findings

Most available information on types of hazardous material

activities and the contaminants associated with these

activities is reliable because it is based on extensive sitespecific data from Federal agency hazardous waste site

cleanup programs. A limitation is that inventories identifying all hazardous waste sites on or adjacent to public land

are incomplete. During field visits, areas of stressed

vegetation, discolored or stained soil and water, indications

of former structures or operations, and land disturbances

may indicate the presence of old, abandoned, or illegal

waste disposal sites. Due to the potentially hazardous nature

of these disposal sites, discovery of such conditions should

be reported to the appropriate agency official for further

action (U.S. Department of Agriculture, Forest Service

1999). Other potential sources of ground water and surface

water contamination, which should be considered in

conducting source water assessments, are cemeteries and

small airports and airstrips, especially those used for aerial

application of chemicals.

Because the need for landfills exists, the design and management of safe landfills are paramount. To meet this need,

Federal, State, tribal, and local governments have adapted

an integrative approach that involves three waste management techniques: (1) decreasing the amount of waste

through source reduction, (2) recycling of materials, and

(3) improving design and management of landfills (U.S.

EPA 1993). A number of regulations exist for the management of a municipal solid waste landfill, and many regulations have flexibility to meet local conditions; managers are

advised to contact a local EPA or State agency office for

information on siting, designing, and managing for their

landfill.

Research Needs

Cleanup of ground water and soil contaminated by solid and

hazardous wastes can be expensive and take long periods of

time. Research is needed to develop more effective and

cheaper cleanup methods.

Key Points

Several factors need to be considered when resource

managers address the effects of landfills on drinking water

quality:

1. Identification of landfill sites—proximity to wells,

aquifers, geological and hydrological features, and

surface waters.

2. Knowledge of the landfill age (a) old landfills—landfills

existing before RCRA may contain hazardous material

and may be improperly designed for hazardous material

storage and municipal waste; (b) existing landfills—

landfills existing after RCRA may still pose a problem for

ground water contamination because the site may contain

older units where hazardous waste was deposited

improperly (these sites may have been improperly

designed or may have punctured liners or clay layers);

and (c) new landfills—landfills being managed under

current Federal and State regulation should pose fewer

problems, but small landfills may be exempted from

certain regulations.

3. Knowledge of landfill history—What was deposited on the

site and when? How was the landfill constructed? Does it

have a clay layer, a liner, or a combination of the two?

4. Monitoring data—Is the site being monitored for VOC’s

and ground water contamination? Is monitoring sufficient

to safeguard ground water sources?

5. Extent of contamination plume—If ground water is

contaminated, what is the vertical and horizontal extent

of the contamination? What is the effect of the plume on

drinking water sources?

6. Compliance with current Federal and State regulations—

What mitigation actions have been taken to comply with

Federal and State laws if contamination occurred?

Chapter 6
72

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Chapter 6
74

Chapter 7

Concentrated Recreation

Myriam Ibarra and Wayne C. Zipperer1

Introduction

This chapter specifically examines drinking water issues

related to concentrated recreation. The effects of concentrated recreation on drinking water quality encompass many

topics with limited published literature. Because of limited

space, selected components—campgrounds, ski resorts,

water recreation, and traffic—are discussed in this report.

However, land managers need to realize that these selected

topics do not include all effects caused by concentrated

recreation on drinking water and serve only to illustrate

potential effects.

One of the most important attractions for public recreation

is public land with natural cover. Increased demands for

outdoor recreation result in greater needs for drinking water

and in increased amounts of wastewater. Expanding

recreation resorts invites larger numbers of visitors, and

private tracts adjacent to public land are magnets for real

estate development. This development may negatively affect

drinking water and alter hydrologic processes. To illustrate

the effects of concentrated recreation on drinking water

supplies, we use data from the Forest Service, but our

findings and recommendations are applicable to managers

of other public and private land.

The National Forest System is the single largest supplier of

public outdoor recreation in the United States. The national

forests offer visitors 4,385 miles of national wild and scenic

rivers; one-third of the National Wilderness Area System;

about 8,000 miles of scenic byways; 133,000 miles of trails;

more than 18,000 campgrounds, picnic areas, and visitor

facilities; and 2.3 million acres of fishing lakes, ponds, and

reservoirs. The Forest Service manages over 23,000

developed facilities, including campgrounds, trailheads, boat

ramps, picnic areas, and visitor centers, in addition to

permitted, privately owned facilities. These facilities can

accommodate approximately 2.1 million people at one time.

In 1997, the Forest Service hosted more than 800 million

recreational visits that included skiing, hiking, camping,

boating, fishing, hunting, and pleasure driving. The number

is expected to grow to 1.2 billion by 2050.

The Forest Service manages over 3,000 drinking water

systems. These systems range in complexity from hand

pump wells to full water treatment plants at major installations. Primarily, these systems use ground water to provide

drinking water at recreation sites and facilities. The Forest

Service manages all public water systems in accordance

with EPA and respective State regulations. In many cases,

this approach exceeds minimum requirements for system

operation.

The principal sources of pollutants produced by concentrated recreation are: (1) fuel residues from automobiles,

watercraft, snowmobiles, and snow making machines;

(2) wastewater from service facilities such as toilets,

showers, restaurants, laundries, etc; and (3) soil and construction materials carried to surface waters with runoff at

the time of construction. Detrimental effects of concentrated

recreation are likely to be episodic or seasonal. The negative

impacts of increased vehicular traffic and concentrated

water recreation may be more apparent on surface water

supplies, while the greater impact of concentrated winter

recreation may be in ground water. This chapter deals with

the effects of increased vehicular traffic, water recreation,

and winter recreation.

Campgrounds

Issues and Risks

The effects of concentrated camping on drinking water

quality are similar to those reported in chapter 8 for dispersed recreation. However, the magnitude, severity, and

frequency of disturbance are greater with concentrated

camping and the associated showers and toilets than with

dispersed camping because of the greater density of humans

using the site. Like other developments, the effects of a

campground on drinking water quality depend on soil

conditions, the presence of vegetation, and existing infrastructure.

1 Hydrologist, 6655 Canton Street, Warner, NY; and Research Forester,

USDA Forest Service, Northeastern Research Station, Syracuse, NY,

respectively.
75

Findings

With the intense use of a site for camping, soil conditions

become extremely important. Soils may lose their organic

layer, become compacted, and become more erodable.

Consequently, more surface erosion may occur as runoff

increases. Without treatment to mitigate effects, the increase

in erosion may result in increased stream turbidity and

sedimentation. Techniques used to minimize soil compaction in urban parks (Craul 1992) may be applicable to

campgrounds. Concentrated camping also could lead to

streambank destabilization and further erosion and sedimentation. The proximity of campgrounds and picnic areas to

water increases the chance of streambank erosion and

destabilization as people use the water for swimming,

bathing, and cleaning cooking and eating utensils.

Vegetation plays a key role in minimizing site degradation.

Vegetation reduces erosion by slowing the movement of

water across the ground surface and increases infiltration of

water by decreasing soil compaction. However, with

increased recreational use, vegetation presence decreases if

active management does not occur to promote vegetation

growth and reduce soil compaction (Craul 1992).

Unlike dispersed camping, concentrated campgrounds

require infrastructure, including parking areas, restrooms,

and shower facilities. This infrastructure may contribute to

the contamination of surface water and ground water. Proper

planning, design, and maintenance of facilities can minimize

contamination of drinking water sources.

Contaminants associated with campers include fecal

material, household cleansers and detergents, garbage and

other floatables, cooking grease and oil, and antifreeze and

motor oil. Because of their remote locations, campgrounds

may serve as sites for illegal dumping of hazardous materials. Enforcement of clean water policies and educational

programs may reduce the levels of these contaminants.

Water Recreation

Issues and Risks

Concentrated recreation on surface water produces chemical

and microbial contamination. Individual boats, marinas, and

swimmers usually release only small amounts of pollutants

that can go undetected. When the number of participants is

large, however, these sources can cause tangible waterquality problems in lakes, reservoirs, and rivers. Boating

and marinas are associated with increased chemical pollutant concentrations and high levels of pathogens in the water

(Gelt 1995, 1998). The effects of swimmers on drinking

water supplies are an emerging problem that has prompted

some utilities to limit or ban recreation on reservoirs used as

drinking water sources. People with weak immune systems

are particularly at risk because current methods for drinking

water treatment do not detect or eliminate all pathogens, and

some residues of chlorination are toxic.

Findings

The use of gasoline with methyl tertiary butyl ester (MTBE)

in motorboats, particularly those using older two-cycle

engines, contaminates surface water (U.S. EPA 1998a). An

estimated 345 million motor boating trips and 29 million jet

skiing trips occurred in the United States during 1994–95

(Cordell and others 1997). Nearly all personal watercraft

and outboard motors use two-cycle engines. The fuelinefficient design of two-cycle outboard motors is essentially unchanged since the 1930’s. Up to 30 percent of the

gas used in the motor goes into the water unburned. Similarly, 10 percent of the fuel used by a personal watercraft,

such as a Jet Ski, leaks into the water.

To assess the impact of two-cycle motorboat engines on

water quality and aquatic life, scientists measured fuel

residues in water in the Lake Tahoe Basin. They found

MTBE; benzene, toluene, ethylbenzene, and xylene

(BTEX); and polycyclic aromatic hydrocarbons (PAH’s)

near shore in lakes that allow motorized watercraft. In open

water, the concentrations of MTBE and BTEX were at or

under the analytical detection limit. On sites with 50 to 100

watercraft engines, MTBE and benzene exceeded drinking

water standards, but concentrations did not approach the

criteria for protection of aquatic life. Concentrations

decreased by the end of the boating season (Allen and

others 1998).

Inefficient two-stroke carburetor engines used in personal

watercraft and as outboard motors are the main source of

fuel pollutants. These engines emitted more than 90 percent

of the MTBE, 70 percent of benzene, and 80 percent of

toluene into Lake Tahoe. In contrast, four-stroke inboard

fuel-injected engines emitted an estimated 8 percent of

MTBE, 28 percent of benzene, and 17 percent of toluene.

Estimated volume of constituent load for Lake Tahoe during

the 1998 boating season from two-stroke engines was in the

order of thousands of gallons of MTBE, hundreds of gallons

of benzene, and tens of hundreds of gallons of tolulene.

There was no evidence that MTBE or BTEX were transported to the bottom of the lake or accumulated there (Allen

and others 1998). Laboratory testing of newer engine

technology suggested that emissions from marine outboard

engines could be virtually eliminated by using more

efficient Ficht injected engines (Allen and others 1998).

Chapter 7
76

Proposed legislation in California moves the implementation

date for stricter EPA emissions controls on personal watercraft engines up 5 years to 2001 from 2006.

Because marinas are located at the water’s edge, pollutants

can go directly to waterways. Water pollution from boating

and marinas is linked to several sources. They include leaks

from underground storage tanks, watercraft engines, and

boat maintenance garages; discharge of sewage from boats;

and stormwater runoff from parking lots (U.S. EPA 1993).

Moreover, physical alteration of shorelines, wetlands, and

aquatic habitat during the construction and operation of

marinas may change flow patterns and result in poorly

flushed waterways.

During boat maintenance, significant amounts of solvent,

paint, oil, and other pollutants potentially can seep into

ground water or be washed directly into surface water.

Paints used to protect boats generally contain toxins that

limit aquatic organism growth. Many boat cleaners contain

chlorine, ammonia, and phosphates that harm plankton and

fish. Small oil spills released from motors and refueling

activities contain petroleum hydrocarbons that may attach to

sediments. Hydrocarbons persist in aquatic ecosystems and

harm bottom-dwelling organisms that are at the base of the

aquatic food web. The EPA recommends that boaters use

nontoxic cleaning products to reduce pollution. Boat owners

can prevent pollution from paint and other chemicals by

vacuuming up loose paint chips and paint dust and by using

a drop cloth when cleaning and maintaining boats away

from the water. Carefully fueling boat engines, recycling

used oil, and discarding worn motor parts into proper

receptacles can prevent needless petroleum spills. Most

importantly, good engine maintenance prevents fuel and

lubricant leaks and improves fuel efficiency (U.S. EPA

1993). Pollution from boating can potentially impair

drinking water reservoirs or seep into ground water wells

that provide drinking water along the shoreline.

Discharge of sewage and waste from boats can degrade

water quality, especially in marinas with high boat use.

Improper disposal of human and pet waste may introduce

pathogenic bacteria, protozoans, and viruses into water (Gelt

1995, U.S. EPA 1993). Sewage from boats can make water

unsuitable for recreation, destroy shell fishing areas, and

cause severe human health problems. Sewage discharged

from boats also stimulates algal growth, which can reduce

the available oxygen needed by fish and other organisms.

Although fish parts are biodegradable, large amounts of

fish-cleaning remains can reduce water quality. Marinas

should have adequate wastewater-disposal hook-ups and

disposal sites for solid waste from boats. Well kept toilet

facilities, designated pet areas, and health education

postings also promote public health.

The locating and design of marinas are two of the most

significant factors impacting water quality. Mastran and

others (1994) found that inlets had higher concentrations of

pollutants than the main channel, suggesting that hydrology

plays a role in the distribution of the pollutants. Poorly

placed marinas disrupt natural water flushing and cause

shoreline soil erosion, habitat destruction, and consequently,

degradation of water quality. Marinas should be located and

designed to be regularly flushed by natural currents. Good

design of a marina can provide an optimum combination of

capacity, services, and access, while minimizing environmental impacts and onsite development costs (U.S. EPA

1993).

Concentrated swimming may cause microbial contamination

of drinking water sources. A study conducted for the

metropolitan water district of southern California determined that a swimmer or bather releases 0.1 gram of feces

when entering the water; infants can add significantly more.

Human feces may harbor viruses, bacteria, protozoa, and

worm pathogens, some of which have been found in water

treated by standard water purification methods. Bacteria are

generally removed by present water treatments. Some

viruses, like hepatitis A and Norwalk, are hardier and can be

controlled only with additional amounts of disinfectant. See

chapter 2 for further discussion on waterborne pathogens.

Water that is accidentally drunk while wading or swimming

poses serious risks. Even small numbers of microbes may

cause disease. It is estimated that in one outing a swimmer

or wader ingests from 0.3 to 1.7 ounces of water that may be

contaminated with feces (Gelt 1998). Outbreaks of

Cryptosporidiosis have been documented from lakes,

community and hotel pools, a large recreational water park,

a wave pool, and a water slide. From January 1995 to

December 1996, 37 outbreaks in 17 States were attributed to

recreational water exposure. Diseases caused by Escherichia

coli O157:H7, a specific strain of E. coli that is known to

cause death if ingested, were associated primarily with

recreational lake water. Cryptosporidium spp. and Giardia

spp. were associated with a few outbreaks in swimming

pools. Outbreaks of Cryptosporidium affected almost 10,000

people, and occurred in swimming pools that were chlorinated (Levy and others 1998).

It is difficult to estimate how many people become sick after

contact with fecal contaminated water. For most people the

symptoms are not acute. A person experiencing diarrhea,

fever, vomiting, and nausea for 2 or 3 days may assume that

he or she has the flu or ate some unsuitable food. In fact, a

Concentrated Recreation
77

person may have a gastrointestinal infection acquired from

drinking water (Gelt 1998). Epidemic outbreaks of waterborne disease have been recognized only after thousands of

acute cases were reported (Levy and others 1998). Isolated

and chronic waterborne diseases probably go undetected or

unrecognized (also see chapter 6 on wastewater treatment).

Methods used to detect enteric pathogens are not always

sensitive to low concentrations but very small numbers

of microbes can cause illness (Gelt 1998). Routine microbiological testing may miss transient contamination by

swimmers. Measures that can be taken to minimize fecal

contamination include: (1) providing changing tables for

infants in locker rooms, (2) providing adequate toilet and

hand washing facilities, (3) posting signs against drinking

water or defecating in the water, and (4) recommending

against swimming for children with gastrointestinal

illness. Unfortunately, other mammals defecating in a

waterbody may introduce enteric pathogens (see chapters

14, 15). Hence, fecal contamination cannot be completely

eliminated.

Winter Recreation

Issues and Risks

The increasing public demand for winter sporting opportunities has led to creation and rapid expansion of skiing resorts

in forested watersheds (Brooke 1999). These facilities may

alter the water quality of pristine environments. The

National Ski Area Association estimates that 60 percent of

all downhill skiing in the United States occurs on national

forests. In cooperation with the 135 ski area operators,

through the National Winter Sports Program, the national

forests provided downhill skiing opportunities to approximately 31 million people in fiscal year 1997. The ski

industry hopes to extend the ski season or even have the ski

resorts open year-round (Hoffman 1998). Some ski resorts

are proposing to develop facilities for summer outdoor

recreation activities such as golf, swimming, and tennis.

With ski resort expansion, real estate development also

expands. To maintain predictable revenues in spite of

unpredictable weather, ski resorts increasingly rely on

artificial snow to cover the slopes. While there is not an

apparent direct effect of skiing on drinking water, environmentalists warn that large ski resorts alter natural hydrological cycles, increase traffic congestion, and are magnets for

urban sprawl, all of which may impair water quality.

Findings

To satisfy public demand, the Forest Service is authorizing

the development or expansion of ski resorts. For example,

between January 1997 and January 1999, the EPA Office of

Federal Activities filed environmental impact statements for

work on 12 ski resorts inside of national forests. Development of ski resorts includes new construction or expansion

of parking lots and service roads, downhill ski runs, crosscountry ski trails, snowmobile trails, chair lifts, lodges,

restrooms, ski patrol facilities, ski schools, ski repair shops,

stores, hotels, and restaurants (U.S. Department of Agriculture, Forest Service 1992, 1998, 1999). The construction and

operation of ski facilities affect drinking water sources in

various degrees. Clearing of vegetation for ski runs increases the chances of soil erosion and hence higher

turbidity and sedimentation in streams (Hoffman 1998).

Pollutants from car emissions are deposited on the soil with

precipitation. Runoff from roads, parking lots, or lawns may

be contaminated with salt, heavy metals, petroleum residues,

or landscaping chemicals. Expansion of impervious surfaces

leads to increased peak runoff and shorter resident time of

water in the watershed.

Newly developed ski resorts may cause shortages or

dramatic fluctuations in drinking water supplies. Some

resorts are projecting to host 5,000 to 10,000 visitors a day.

The typical average consumption rate of water at ski areas is

10 gallons per day per skier capacity; if water conservation

measures are in place, the intake could be reduced to 7

gallons per day. Thus, a ski resort with 13,000 skiers may

need between 94,500 and 135,000 gallons per day (U.S.

Department of Agriculture, Forest Service 1998). At the

same time, a small but irretrievable loss of ground water

may occur due to evaporation and sublimation from snow

making (Hoffman 1998). To prevent artificially drastic

pulses in downstream flow and to maintain channel stability,

ski resorts may need to stop making snow when natural

water levels are too low, or use water stored in ponds or

lakes (U.S. Department of Agriculture, Forest Service

1997).

Ski resorts are often located on environmentally sensitive

sites. In mountainous regions, the slopes are steep, the soils

are thin, the subsurface is predominantly gravel and cobble,

and the aquifers are fractured bedrock. This type of aquifer

is very sensitive to pollution because the rapid ground water

flow can carry microbes and other pollutants for long

distances (U.S. EPA 1999). Ski resorts have a special

problem with wastewater treatment. The peak need is in the

winter, when conventional sewage treatment methods

function at slower rates and microbial pathogens survive

longer in water and soil. One solution is to build storage

ponds and apply wastewater treatment in warmer weather.

Such storage, however, is not always economically or

logistically feasible. Another method being tested makes

artificial snow from wastewater and stores the snow on

Chapter 7
78

slopes where skiing is not permitted (Gibson 1996). In ideal

conditions, the wastewater stored would melt and percolate

very slowly, producing a clean effluent. However, sudden

snowmelt could contaminate surface water and ground

water with effluent.

Ski resorts rely more and more on snow making and

grooming to attract skiers. Some snow making operations

require massive amounts of stream water. To get enough

water, resorts have relocated stream channels, excavated

wells, constructed ponds or pumped water from neighboring

surface water sources. Each activity may alter the natural

flow of water and ultimately influence drinking water

quality. Not only is water being redistributed to another

location, the generators that power the snow machines and

pumps may contribute to air pollution. For instance, the

diesel generators in one resort in Vermont are the eighth

largest air polluters in the State (Hoffman 1998). This

pollution may contribute to atmospheric deposition of

contaminants.

Increased Traffic

Issues and Risks

Vehicular traffic in forests and grasslands creates fuel

emissions that are deposited on the ground through wet and

dry deposition. Pollution from fuel emissions may migrate

to surface water and ground water through rain or snowfall.

The most significant sources of fuel pollutants are cars, but

in some places all-terrain vehicles and snowmobiles also are

important contributors. In the last decade, the number of

recreational visits to national forests increased by 40

percent, and the number of visits was highly correlated with

the number of vehicles (Cordell and others 1997). Additionally, tourism to recreational resorts promotes urbanization,

which in turn adds traffic. For example, in the Eisenhower

Tunnel connecting Denver with the busiest ski areas in

Colorado, the traffic has quadrupled in the last 25 years.

Improvement and expansion of parking lots and roads

increase peak runoff and nonpoint-source pollution from

impervious surfaces. Runoff may be contaminated with salt,

heavy metals, petroleum residues, or landscaping chemicals

that can degrade surface water and ground water quality

(U.S. EPA 1983). Oxygenates and PAH’s are gasoline

residues that have been found in drinking water supplies and

are potential threats to human health. Deposition of MTBE,

an oxygenate, may be especially significant during the

winter because concentrations of MTBE in precipitation are

higher at colder than warmer temperatures (Delzer and

others 1996). Widespread impacts may result from vehicular

emissions that dissolve in rain or snowfall and subsequently

infiltrate to shallow ground water. Additional research needs

to be conducted to determine the significance of increased

auto emissions on drinking water quality in rural areas.

Auto emissions also contribute to the amount of nitrogen in

the atmosphere. Nitrogen deposition from the atmosphere

varies across the country with the greatest concentrations

occurring in a broad band from the Upper Midwest through

the Northeast (U.S. Geological Survey 1999). Recent studies

have shown that atmospheric deposition of nitrogen can be

quite significant. For example, approximately 25 percent of

the nitrogen entering the Chesapeake Bay Estuary comes

from the atmosphere (Fisher and Oppenheimer 1991). The

effect of nitrogen deposition on drinking water is an area

that needs further research (see chapter 3). See chapter 6 for

further discussion of urban runoff and MTBE.

Reliability and Limitation of Findings

The potential negative impacts of concentrated recreation on

drinking water supplies have been recognized and addressed

in a qualitative way, but quantitative assessments are very

rare. The material presented here comes almost exclusively

from government reports and newspaper articles rather than

from the primary scientific literature. This fact suggests that

the issue has not been subjected to adequate scientific

investigation.

A simplistic first approximation is to consider the expansion

of concentrated recreation in forests as small-scale urbanization. However, it is important to keep in mind that the

toxicities of some pollutants produced by recreational

activities have been measured only in the laboratory.

Furthermore, survey data on impacts on water quality by

recreation are mostly from water that is not used for

drinking (Cox 1986, Gelt 1995). The extremely varied

ecology of each forest together with the diverse nature of

recreation activities suggests specific analysis for each

situation. Drinking water sources seldomly appear to be

susceptible to long-term degradation because of recreation,

but some lakes and well water probably are susceptible to

episodes of local pollution (Peavy and Matney 1977).

Environmental impact statements are prepared when

designing recreation resorts, and they often present plans to

monitor surface water and ground water. In the absence of

specific studies, analysis of these data could be the first step

in describing regional or national patterns.

The EPA and Centers for Disease Control and Prevention

(CDC) recognized that waterborne diseases are common in

the United States (Levy and others 1998), but data on their

occurrence are very sparse. The United States has recorded

incidences of waterborne diseases only since 1985. The EPA

Concentrated Recreation
79

and CDC are conducting a series of pilot studies to produce

the first national estimate of waterborne disease occurrence

(U.S. EPA 1998b). Of particular importance are the levels of

disease associated with drinking water that otherwise meet

Federal and State standards. This research would also serve

as a springboard for more localized assessments of drinking

water quality.

General principles of urban water pollution are applicable in

expanding recreation resorts in all regions. However, data to

quantify impacts on specific sites are not readily available.

The Forest Service and others have monitoring programs

that document some aspects of water resources, but we are

not aware of any efforts to collect data specifically to

evaluate the impact of concentrated recreation on drinking

water supplies.

Research Needs

The field of recreation ecology is relatively new. Only

recently have scientists begun to study the relationships

among use-related, environmental, and managerial factors

(Marion 1998). Evaluation of the effect of recreation on

drinking water could be approached through monitoring the

effects of the visitor population and the impacts of population growth in communities adjacent to recreation sites.

1. One basic task is to document the kinds of data that have

been collected as part of routine water-quality monitoring

and sanitary engineering operations.

2. The next step is to design a sampling program for

evaluating impacts of recreation on drinking water

supplies.

3. Impacts also need to be assessed across the range of scale

from local to major watershed.

Key Points

1. Concentrated recreation, like urbanization, affects water

quality through wastewater treatment and urban runoff.

2. Ski resorts alter hydrologic processes by changing the

availability of water during the year.

3. Decreased streamflows may increase the concentration of

contaminants from wastewater and runoff.

4. Wastewater treatment is especially precarious for ski

resorts because peak treatment is during the winter.

5. Water recreation, both swimming and boating, may have

direct effects on drinking water quality at the local scale.

6. Increased traffic may affect drinking water quality

through deposition of MTBE and nitrogen.

Literature Cited

Allen, J.; Reuter, J.; Goldman, C. [and others]. 1998. Lake Tahoe

motorized watercraft report—Integration of water quality, watercraft use

and ecotoxicology issues. Davis, CA; University of California; Reno,

NV: University of Nevada; preliminary draft report. 24 p.

Brooke, J. 1999. Environmentalists battle growth of ski resorts. The New

York Times. January 19: Sec. A; 10 (col. A1).

Cordell, H.K.; Teasley, J.; Super, G. [and others]. 1997. Results from the

national survey on recreation and the environment, all Forest Service

regions. Athens, GA: U.S. Department of Agriculture, Forest Service;

University of Georgia. [Available on the Web at: http:/www.fs.fed.us/

research/rvur/recreation/].

Cox, E.R. 1986. Wastewater movement near four treatment and disposal

sites in Yellowstone National Park, Wyoming. WRI–84–4356. Denver:

U.S. Geological Survey. 81 p.

Craul, P.J. 1992. Urban soils in landscape design. New York: John Wiley.

396 p.

Craun, G.F.; Calderon, R. 1996. Microbial risks in groundwater systems,

epidemiology of waterborne outbreaks. In: Under the microscope:

Proceedings of the groundwater foundations’s 12th annual fall symposium; 1996 September 5–6; Boston. Denver: American Water Works

Association Research Foundation: 9–15.

Delzer, G.C.; Zogorski, J.S.; Lopes, T.J. [and others]. 1996. Occurrence

of the gasoline oxygenate MTBE and BTEX compounds in urban

stormwater in the United States, 1991–1995. Rep. 96–4145. Denver:

U.S. Geological Survey, Water-Resources Investigations. 8 p.

Fisher, D.C.; Oppenheimer, M. 1991. Atmospheric nitrogen deposition

and the Chesapeake Bay Estuary. AMBIO. 20(3–4): 102–108.

Gelt, J. 1995. Water recreation makes big splash in Arizona. Arroyo. 7:

1–11.

Gelt, J. 1998. Microbes increasingly viewed as water quality threat, the

emerging contaminants. Arroyo. 10: 1–14.

Gibson, T. 1996. A cool solution for treating wastewater. Compressed Air.

101(8): 18–22.

Hoffman, C. 1998. Let it snow. Smithsonian. 29: 50–53.

Levy, D.A.; Bens, M.S.; Craun, G.F. [and others]. 1998. Surveillance for

waterborne-disease outbreaks—United States, 1995–1996. Morbidity

and Mortality Weekly Report. 47(SS–5): 1–34.

Marion, J.L. 1998. Recreation ecology research findings: implications for

wilderness and park managers. In: Proceedings of the national outdoor

ethics conference; 1998 April 18–21; St. Louis. Gaithersburg, MD: Izaac

Walton League of America: 188–196.

Mastran, T.A.; Dietrich, A.M.; Gallagher, D.L. [and others]. 1994.

Distribution of polyaromatic hydrocarbons in the water and sediments of

a drinking water reservoir with respect to boating activity. Water

Research. 28: 2353–2366.

Peavy, H.S.; Matney, C.E. 1977. The effects of recreation on water quality

and treatability; drinking water quality enhancement through source

protection. In: Pojasek, R.B., ed. Drinking water quality enhancement

through source protection. Ann Arbor, MI: Ann Arbor Science Publishers

Inc.: 461–475.

Chapter 7
80

U.S. Department of Agriculture, Forest Service. 1992. Loon Mountain

ski area, South Mountain expansion project, Grafton County, NH. Final

environmental impact statement. 247 p. On file with: White Mountain

National Forest, Federal Building, 719 Main Street, Laconia, NH

03246–0772.

U.S. Department of Agriculture, Forest Service. 1997. Record of

decision for the Waterville Valley Ski Resort development of

snowmaking water impoundments and environmental impact statement.

39 p. On file with: White Mountain National Forest, Federal Building,

719 Main Street, Laconia, NH 03246–0772.

U.S. Department of Agriculture, Forest Service. 1998. Pelican Butte ski

area master development plan: implementation. WIN–24–98: III–19 to

III–83. On file with: Winema National Forest, 2819 Dahlia, Klamath

Falls, OR 97601.

U.S. Department of Agriculture, Forest Service. 1999. Final environmental impact statement for permitting Appalachian Mountain Club

(AMC) huts and Pinkham Notch Visitor Center (PNVC) in the White

Mountain National Forest. 279 p. On file with: White Mountain National

Forest, Federal Building, 719 Main Street, Laconia, NH 03246–0772.

U.S. EPA. 1983. Results of the nationwide urban runoff program. PB84–

185552. Washington, DC: U.S. Environmental Protection Agency. 188 p.

U.S. EPA. 1993. Guidance specifying management measures for sources of

nonpoint pollution in coastal waters. EPA–840–B–93–001c. Washington,

DC: U.S. Environmental Protection Agency, Office of Water, Nonpoint

Source Pollution Control Program. 828 p.

U.S. EPA. 1998a. Class V injection wells. EPA/816–F–98–009. Washington, DC: U.S. Environmental Protection Agency, Office of Water. 4 p.

U.S. EPA. 1998b. Overview of the Federal UST program. Washington, DC:

U.S. Environmental Protection Agency. [Not paged]. [Available on the

Web at: http://www.epa.gov/swerust1/overview.htm].

U.S. EPA. 1999. Ground water rule proposal. Draft document EPA 99–022.

Washington, DC: U.S. Environmental Protection Agency. 78 p.

U.S. Geological Survey. 1999. The quality of our Nation’s water—

nutrients and pesticides: Circ. 1225. Denver: U.S. Geological Survey.

82 p.

Concentrated Recreation
81

Chapter 8

 Dispersed Recreation

David Cole1

Introduction

Dispersed recreation is a common and growing use of

forests and grasslands that has the potential for significant

impacts on the quality of public drinking water sources.

Issues and Risks

Trails are constructed to provide access. Visitors walk, ride,

and bicycle along trails. Runoff from trails can add sediment

to streams, particularly at trail fords. Visitors picnic, camp,

and walk or ride off-trail; in some places, they use off-road

vehicles to travel cross country. The resultant loss of

vegetation and compaction of soil can lead to increased

runoff, erosion, and sedimentation. Visitors who pull off

roads to view scenery, picnic, camp, or access the immediate

surroundings can cause increased erosion and sedimentation

of streams. Where people pull off roads to picnic or camp

adjacent to streams, foot traffic and vegetation loss on

streambanks can result in streambank erosion and channel

instability.

Visitors and their animals can contaminate water supplies by

carrying and depositing feces containing microorganisms

that cause human diseases. Contamination comes from fecal

deposition and from direct contact with water during

activities, such as swimming and washing. Recreational

behaviors are commonly unrestricted, visitor education is

typically inadequate, and where activities are dispersed, few

facilities are provided to ensure proper disposal of human

waste. Consequently, drinking water quality problems

associated with recreation use may be expected. In a recent

survey of Forest Service watershed managers, recreation

was the most commonly reported cause of water-quality

concerns. However, this high frequency of concern does not

necessarily mean that recreation is the most common or

serious source of water contamination in national forest

watersheds.

Findings from Studies

The impacts of dispersed recreation on sediment have not

been systematically quantified. Recreation facilities (particularly trails) and recreation use elevate sediment levels

(see chapter 9). Nonmotorized recreation simply does not

disturb much of the watershed. Cole (1981) found, for

example, that <0.5 percent of a heavily used portion of the

Eagle Cap Wilderness in Oregon was directly affected by

trails and camping. Most of the disturbed area was located

far enough from streams so that the effect was negligible.

Recent research indicates that sediment yield from trails is

much higher when trails are used by horses than by hikers

or llamas (DeLuca and others 1998).

Impacts of dispersed motorized recreation activities on

sediment, while not well quantified, are more likely to be

significant. Impacts will vary greatly with such factors as

type of vehicle, driving behavior, topography, vegetation

type, soil erodibility, and climate. Both the extensiveness

and the intensiveness of impact are much greater with

motorized recreation than with nonmotorized recreation. In

the extreme case of an off-road vehicle area in California,

erosion rates were estimated to be 52 tons per acre per year

(116.5 metric tonnes per hectare per year) (Wilshire and

others 1978).

Pathogenic organisms can be introduced by recreationists

into watersheds in which dispersed recreation is the primary

land use. In a broad survey of surface municipal drinking

water sources, LeChevallier and others (1991) found oocysts

of Cryptosporidium spp. and cysts of Giardia spp. species

even in protected watersheds. Suk and others (1987) found

cysts of Giardia in 27 of 78 samples from back-country

streams in several large wilderness areas in the Sierra

Nevada in California. Taylor and others (1983) found

Campylobacter jejuni in the stools of 23 percent of people

reporting diarrhea and G. lamblia in the stools of 8 percent

of such people. They also found these organisms in streams

in the Grand Teton National Park, WY.

It is generally accepted, although still controversial, that

mammals other than humans can spread these pathogenic

organisms to humans. Since horses, mules, and dogs are

1

 Research Biologist, Aldo Leopold Wilderness Research Institute,

Missoula, MT.
82

more likely than humans to defecate directly in or near

water, they may be a major concern if they are important

disease carriers. Taylor and others (1983) found

Campylobacteria in a sample of packstock stool in Grand

Teton National Park, as well as in samples from humans.

More than 40 mammals, both wild and domestic, have been

found to harbor Cryptosporidium parvum (Current 1987).

This evidence, along with the finding that C. parvum readily

crosses host species barriers, has convinced most experts

that human infections are often the result of transmission

from wild and domestic animals, including horses and dogs

(Current and Garcia 1991, Rose 1990). As for Giardia,

Hibler and Hancock (1990) state, “some investigators

considered the parasite found in humans (Giardia lamblia)

to be host-specific, but the majority of the research performed to date questions this assumption.” Cryptosporidium

and Giardia have been found in humans and a wide variety

of birds, mammals, fish, and reptiles (see appendix D). It

has been cross transmitted between humans and a number of

these animals (Hibler and Hancock 1990).

The issue of transmission by wild animals (chapter 14) is

also relevant to the question of whether or not water quality

can be adequately protected by eliminating or severely

restricting recreation use. Questions remain about which

pathogens are transmitted and the relative importance of

humans and other animals as agents of transmission.

Consequently, management actions such as the improvement of human waste disposal behavior and facilities, and

even outright elimination of recreation use, while likely to

reduce the transmission of disease organisms, are unlikely to

eliminate the problem.

Studies that have attempted to relate intensity of recreation

use to degree of water contamination have produced mixed

results. Some studies report positive correlations (e.g., Suk

and others 1987), others report no correlation (e.g.,

Silverman and Erman 1979), and at least one series of

studies reports a negative correlation (Stuart and others

1971, Walter and Bottman 1967). One potential explanation

for these divergent findings is that wild animal contamination may dwarf the effects of low levels of recreation.

Indeed, some authors have noted that as levels of contamination increase, the strength of positive correlations between

recreational use and contamination and between fecal

coliform and the occurrence of Giardia and

Cryptosporidium also increase (LeChevallier and others

1991).

The study finding a negative correlation between recreation

use and bacterial contamination of water supplies initially

compared a watershed closed to recreation use with a

watershed open to use. Fecal coliform and fecal streptococci

counts were higher in the closed watershed (Walter and

Bottman 1967). After the watershed was opened to recreation and limited logging, bacterial contamination decreased. They concluded, “...these human activities drove

from the watershed a large wild animal population which

had contributed substantially to the previous bacterial

population” (Stuart and others 1971: 1048).

From these findings, several implications can be drawn.

First, surface water is not likely to be safe for drinking

without purifying treatment, even where recreation use is

excluded. In fact, Suk and others (1987) found in wilderness

watersheds that 45 percent of high-use samples contained

Giardia cysts, and 17 percent of the low-use samples

contained cysts. Back-country visitors are advised to purify

drinking water obtained from all surface water sources,

regardless of the level of recreation use in the vicinity

(Cilimburg and Monz, in press). Adequate purifying

treatment for public drinking water may be expensive.

Second, it is more critical to improve management of

recreation use and of human waste disposal in heavily used

than in lightly used watersheds. Management options for

areas with heavy dispersed recreation use include reducing

recreation use, prohibiting pack animals and pets, providing

adequate toilet facilities, and educating visitors in appropriate waste disposal techniques (see, e.g., Hampton and Cole

1995, Meyer 1994).

The relationship between the amount of dispersed recreation

and water contamination depends on other variables

including the type of recreation use, soils, slope, and

climate. None of these relationships has been systematically

evaluated. It is difficult to determine if recreation use is

heavy or light, or to confidently prescribe management in

field situations.

The importance of educating visitors in the proper disposal

of human waste is suggested by studies of the survival of

bacteria in feces buried in soil in Montana. Samples of feces

were inoculated with two bacteria, Escherichia coli and

Salmonella typhimurium, and both survived in large

numbers for 8 weeks after burial in early summer (Temple

and others 1980). Moreover, substantial numbers of Salmonella survived over winter. Depth of burial had no effect on

persistence, and differences among burial sites were minor

(Temple and others 1982). Clearly, the idea that shallow

burial (in catholes) renders feces harmless in a short time is

inaccurate. Removal of feces is the best means of disposal if

toilets are not provided. The second best option is careful

and complete burial far from water sources, campsites, and

other heavily visited locations.

Dispersed Recreation
83

Reliability and Limitation of Findings

There is strong evidence to support the general findings that

(1) dispersed recreation use can adversely affect the quality

of surface drinking water supplies and (2) surface drinking

water supplies will contain pathogenic microorganisms even

in the absence of recreation use. Our ability to quantify the

effect of dispersed recreation is very limited, as is our

understanding of the importance of recreation as a source of

contamination. Consequently, there is a weak foundation in

science for decisions about where recreation use should be

prohibited or restricted and where sanitary facilities should

be provided or improved.

These general findings should be broadly applicable

throughout the United States. Specifics of quantitative

relationships between recreation use and water quality will

vary with many environmental parameters. Logic suggests

that one important regional distinction can be made between

arid and mesic regions. In arid lands, visitors and animals

are particularly drawn to water sources, increasing the

likelihood of contamination and the effects may persist

longer because these systems are not flushed rapidly or

frequently.

Research Needs

1. We need to know if some pathogens, such as human

enteric viruses, pose a significant threat to human health.

As Gerba and Rose (1990) note, even though there are

few cases where virus isolations in source water have

been linked to human disease, there are many reasons to

suppose that there is much more illness due to viral

contamination than is recognized. We need a better

understanding about the mechanisms of transmission for

different pathogenic microorganisms, especially their

presence in recreation pack animals, pets, and wild

animals. Further research on C. parvum and G. lamblia is

particularly important.

2. Additional research is needed to provide a more solid

foundation for decisions about where and how to restrict

dispersed recreation and where to invest in more and

better sanitary facilities. We need better quantification of

the relationship between drinking water microbiology

and amount of use by visitors, their pets, and their pack

animals. Thresholds of use need to be identified, above

which adverse effects on water quality become pronounced and unacceptable. We need a better understanding of how site variables influence susceptibility to

contamination and whether water-contact activities, such

as swimming, are a significant concern at the low

densities typical of dispersed recreation sites.

3. Research is needed to develop techniques capable of

distinguishing between human and other sources of

pathogens. Finally, we should assess (1) the validity of

rules of thumb managers use to develop management

prescriptions and (2) the effectiveness of techniques

managers develop to mitigate contamination.

4. More research is needed on the decomposition rates of

human feces, on variables that influence decomposition

rates, and on how pathogens disperse in and over the soil.

This information could contribute to better educational

material about where and how to bury feces, and to better

decisions about where sanitary facilities are needed.

Key Points

1. Since dispersed recreation can contribute to contamination, every affordable effort should be made to educate

visitors in appropriate human waste disposal and to

provide well-designed and appropriately located facilities

for the disposal of human waste. Surface water from

wildlands including wilderness, can contain pathogens

that cause human disease unless drinking water is

adequately treated. For public water supplies, adequate

treatment may be expensive.

2. Where recreation use is high and water contamination is

too high, sanitary facilities need to be developed or

improved, and/or use of the area must be restricted.

Where it is clear that dispersed recreation use is low, use

restrictions and the provision of sanitary facilities are not

worth the costs involved. In areas of moderate use, our

understanding is inadequate to suggest whether it is

worth the costs of limiting access, restricting behavior, or

investing in sanitary facilities. Inadequate understanding

also makes it difficult to identify use thresholds above

which (1) rudimentary sanitary facilities are needed or

(2) developed sanitary facilities are needed.

Chapter 8
84

Literature Cited

Cilimburg, Amy; Monz, Christopher [In press]. Wildland recreation and

human waste: a review of problems, practices and concerns.

Environmental Management.

Cole, David N. 1981. Vegetational changes associated with recreational use

and fire suppression in the Eagle Cap Wilderness, Oregon: some

management implications. Biological Conservation. 20: 247–270.

Current, William L. 1987. Cryptosporidium: its biology and potential for

environmental transmission. Critical Reviews Environmental Control.

17: 21–51.

Current, William L.; Garcia, Lynne S. 1991. Cryptosporidiosis. Clinical

Microbiology Reviews. 4: 325–358.

DeLuca, T.H.; Patterson, W.A.; Freimund, W.A.; Cole, D.N. 1998.

Influence of llamas, horses,and hikers on soil erosion from established

recreation trails in western Montana, U.S.A. Environmental Management. 22: 255–262.

Gerba, Charles C.; Rose, Joan B. 1990. Viruses in source and drinking

water. In: McFeters, Gordon A., ed. Drinking water microbiology. New

York: Springer-Verlag: 80–396.

Hampton, Bruce; Cole, David. 1995. Soft paths. 2d

 ed. Mechanicsburg,

PA: Stackpole Books. 222 p.

Hibler, Charles P.; Hancock, Carrie M. 1990. Waterborne giardiasis. In:

McFeters, Gordon A., ed. Drinking water microbiology. New York:

Springer-Verlag: 271–293.

LeChevallier, M.W.; Norton, W.D.; Lee, R.G. 1991. Occurrence of

Giardia and Cryptosporidium spp. in surface water supplies. Applied

Environmental Microbiology. 57: 2610–2616.

Meyer, Kathleen. 1994. How to shit in the woods. Berkeley, CA: Ten

Speed Press. 77 p.

Rose, Joan B. 1990. Occurrence and control of Cryptosporidium in

drinking water. In: McFeters, Gordon A., ed. Drinking water microbiology. New York: Springer-Verlag: 294–321.

Silverman, G.; Erman, D.C. 1979. Alpine lakes in Kings Canyon National

Park, California: baseline conditions and possible effects of visitor use.

Journal of Environmental Management. 8: 73–87.

Skinner, Q.D.; Adams, J.C.; Rechard, P.A.; Beetle, A.A. 1974. Effect of

summer use of amountain watershed on bacterial quality. Journal of

Environmental Management. 3: 329–335.

Stuart, D.G.; Bissonnette, G.K.; Goodrich, T.D.; Water, W.G. 1971.

Effects of multiple use on water quality of high-mountain watersheds:

bacteriological investigations of mountain streams. Applied Microbiology. 22: 1048–1054.

Suk, Thomas J.; Sorenson, Stephen K.; Dileanis, Peter D. 1987. The

relation between human presence and occurrence of Giardia cysts in

streams in the Sierra Nevada, California. Journal of Freshwater Ecology.

4: 71–75.

Taylor, D.N.; McDermott, K.T.; Little, J.R. [and others]. 1983.

Campylobacter enteritis from untreated water in the Rocky Mountains.

Annals of Internal Medicine. 99: 38–40.

Temple, Kenneth L.; Camper, Anne K.; Lucas, Robert C. 1982.

Potential health hazard from human wastes in wilderness. Journal of Soil

and Water Conservation. 37: 357–359.

Temple, Kenneth L.; Camper, Anne K.; McFeters, Gordon A. 1980.

Survival of two enterobacteria in feces buried in soil under field

conditions. Applied and Environmental Microbiology. 40: 794–797.

Walter, W.G.; Bottman, R.P. 1967. Microbiological and chemical studies

of an open and closed watershed. Journal of Environmental Health. 30:

159–163.

Wilshire, H.G.; Nakata, J.K.; Shipley, S.; Prestegaard, K. 1978. Impacts

of vehicles on natural terrain at seven sites in the San Francisco Bay

area. Environmental Geology. 2: 295–319.

Dispersed Recreation
85

Chapter 9

Roads and Other Corridors

W.J. Elliot1

Introduction

The focus of this chapter is on the impact on drinking water

quality of roads and other corridors, such as trails, utility

rights-of-way, railroads, and airfields in forest and grassland

watersheds. These corridors are essential for a wide range of

access including residential, recreational, and managerial

(U.S. Department of Agriculture, Forest Service 1998).

They can be public and managed by Federal, State, or

local agencies or private and managed by individuals or

industries.

Roads, railroads, and similar corridors are major features in

most watersheds. Figure 9.1 is a diagram of a typical

insloping forest road in steep terrain. Water from the

roadway is diverted to the ditch, and then directed to a

culvert or surface drain. In less steep areas, or for larger

roads, there are usually ditches on both sides of the road to

collect and channel runoff. The runoff is then delivered to

vegetated slopes for infiltration or to a natural channel that

is part of the stream system (Packer and Christensen 1977).

Roads and similar corridors can adversely affect water

quantity and quality in several ways. Runoff is low from

undisturbed forests, but runoff rates from rainfall and

snowmelt are greater from compacted road surfaces than

from less disturbed parts of watersheds (Elliot and Hall

1997). The roadway, the ditch, and in some cases, the

waterway below a road culvert are the main sources of

detached sediment (fig. 9.1) from erosion depending on road

surface material (Elliot and Tysdal 1999). The cutslopes and

fillslopes erode mainly by mass wastage.

Eroded sediment is usually deposited on the undisturbed

surface below the road (Elliot and Tysdal 1999, McNulty

and others 1995, Packer and Christensen 1977). Establishing

a buffer zone of undisturbed forest between a corridor and a

stream is helpful, but if runoff from roads or other disturbances is channeled, or filter strips are too narrow, then

buffer zones cannot be expected to eliminate sediment

movement to streams. Most surface water contaminants

enter streams at stream crossing by roads, railroads, or

pipelines, or places where other disturbances are close to

streams. Corridor-related disturbances also can degrade

ground water from shallow wells, particularly in highly

porous geologies, such as karst (Gilson and others 1994,

Hubbard and Balfour 1993, Keith 1996).

Excavation at the bottom of a cutslope can intercept ground

water, creating instability of the road or the cutslope and

altering hydrology (Jones and others, in press). This

intercepted ground water may also be affected by acid

drainage (chapter 18). All of the excavated surfaces revegetate slowly and are prone to erosion (Burroughs and King

1989, Grace and others 1998).

When roads or other compacted corridors are abandoned,

they can continue to be sources of sediment through chronic

surface erosion or mass failure (Elliot and others 1996).

Compaction of a disturbed surface frequently restricts

vegetation regrowth. Bare surfaces are susceptible to

erosion, and steep areas without trees are susceptible to

landslides. In some cases, local frost heaving or minor

slumping of fill shoulders can cause surface water to collect,

leading to saturation of the fill and an increased risk of mass

failure. Both surface erosion and mass failure can lead to

1

 Project Leader, USDA Forest Service, Rocky Mountain Research Station,

Moscow, ID. Figure 9.1—Surfaces and flow paths associated with a road cross section.

Channel

Plan

Culvert

Section

 Roadway

Cutslope

Ditch

Fillslope

Original surface

Original surface

Water table
86

Roads and Other Corridors

increased sediment loads in streams (McClelland and others

1998).

Runoff and seepage from roads and rights-of-way can

contain elevated levels of sediment, metals, and complex

hydrocarbons from the highway material and traffic. They

may also contain traces of pesticides or other undesirable

substances. Chemicals may be dissolved in the runoff water,

but they frequently are attached to the eroded sediment

particles.

Altered Hydrology

Issues and Risks

The presence of roads in a watershed may increase the

frequency and magnitude of peak runoff discharges,

particularly on small watersheds. Roads may also increase

total runoff and decrease the time to peak runoff from major

storms or snowmelt.

Findings from Studies

Roads have a number of impacts on hydrology. They

intercept precipitation and snowmelt and, because they have

lower infiltration rates, divert it as surface runoff to channels

(Packer and Christensen 1977). Cutslopes (fig. 9.1) can also

tap into ground water and divert it, increasing runoff. In a

study of spring snowmelt in the northern Rockies, 58

percent of the runoff from a road was due to intercepted

subsurface flow (Burroughs and Marsden 1972). Megahan

(1983) found that road segments on granitic soils in central

Idaho collected about 8.4 inches [21 centimeters (cm)] of

water in subsurface flow from the area above the road. Road

ditches can extend the stream network, increasing the

volume of water available during the early part of a storm.

The presence of roads can also shorten the time to peak flow

during a runoff event (Wemple and others 1996). This

diversion of ground water can dry out hillsides below the

road, altering vegetation, and reducing water yield during

dry periods later in the year.

If a road culvert is too small, or becomes blocked, water can

be diverted from one subwatershed to another. Severe ditch

and channel erosion may result (Megahan 1983; U.S.

Department of Agriculture, Forest Service 1998). The

cumulative effect is an increase in frequency and magnitude

of peak discharges (Jones and Grant 1996; Megahan 1983;

U.S. Department of Agriculture, Forest Service 1998).

Adding gravel to the road surface increases the porosity and

roughness of the road, increasing the conductivity from

under 1 millimeter (mm) per hour to 3 mm per hour or more

(Foltz 1996). This results in decreased runoff rates from

low-intensity rain and snowmelt. Gravel addition will have

less impact during high-intensity storms. Ripping closed

roadways can increase infiltration rates, but studies show

that rates do not reach undisturbed levels (Luce 1997).

Culverts or surface drainage structures to deliver water to

hillsides rather than to channels will also reduce the hydrologic impacts of roads (Elliot and Tysdal 1999).

Reliability and Limitations of Findings

Generally, roads will have the same types of impacts on

hydrology regardless of climatic or soil differences, but the

magnitude of impact may vary substantially (Elliot and

others 1999a). Impacts of disturbances and benefits of

mitigation measures will be greater in wetter climates.

Interception of subsurface flows depends on slope position,

depth to the water table, and availability of subsurface flow.

The greatest challenge in applying the hydrologic findings is

that landscapes are highly variable, making differences in

hydrology due to the presence of roads difficult to isolate.

Research Need

The main research need is watershed scale studies to

compare relatively undeveloped watersheds to similar

watersheds with greater disturbances due to roads. Such

sites are difficult to find, so hydrologic predictive models

need to be developed and verified.

Key Point

Roads in a watershed may increase the amount of runoff and

the peak runoff rate.

Sedimentation

Issues and Risks

On most forested watersheds, sediment is the most troublesome pollutant and roads are a major source of that sediment

(Appelboom and others 1998; Megahan and Kidd 1972a,

1972b; Patric 1976; Reid and Dunne 1984; Yoho 1980).

Sediment can adversely impact water quality by increasing

turbidity, prematurely plugging filters and other components

of treatment systems. Suspended sediment can also carry

undesirable chemical pollutants, such as phosphates,

pesticides, and other hydrocarbons into surface water and

ground water (Gilson and others 1994, Patric 1976,

Thomson and others 1997). See chapter 3 for additional

impacts of sediment.
87

Sediment may be from surface erosion, which is generally

more likely to carry pollutants. On steep watersheds, more

sediment may be from mass wasting, which tends to bring

greater volumes of soil to the stream.

Findings from Studies

Numerous researchers and managers throughout the United

States have identified roads as a major source of sediment in

otherwise relatively undisturbed watersheds, such as forests

and rangelands (table 9.1). Table 9.1 presents some typical

erosion rates for different regions in the United States for

different types of disturbance. Note that some investigators

have reported erosion rates for roads, ranging from 5 to 550

tons per acre [11.2 to 1232 metric tonnes (Mg) per hectare]

per year, whereas others have reported erosion rates of

watersheds containing roads in the range of 0.02 to 2 tons

per acre (0.045 to 4.5 Mg per hectare) per year. The wide

range results from differences in measuring erosion (at the

road or at the watershed outlet) and in the factors causing

erosion, including the presence, density, and design of the

road network on the watershed.

In a mixed rural and urban watershed in northern Idaho,

roads covered only 1 percent of a watershed, but they

contributed 8 percent of the sediment to streams (Idaho

Division of Environmental Quality 1997). Megahan (1974)

estimated that, in central Idaho, the sediment yield from

watersheds without roads was about 0.07 tons per square

mile (0.025 Mg per square kilometer) per day, whereas the

presence of roads increased this yield by a factor of 5.

McNulty and others (1995) attributed the majority of

sediment from a forested watershed in the Southeast to

unpaved roads.

Immediately after roads are constructed, erosion rates from

bare slopes and road surfaces are high (fig. 9.2). Erosion

rates can drop rapidly as exposed slopes revegetate and

stabilize. Erosion reductions of 90 percent or more are

common as a road ages (Burroughs and King 1989,

Ketcheson and Megahan 1996). Road surfaces, however,

will likely continue to be a source of sediment as long as

traffic or maintenance prevents the establishment of

vegetation (Elliot and others 1996, Swift 1984b). Applications of high-quality gravel to unpaved roads can decrease

erosion rates by up to 80 percent (fig. 9.2) (Burroughs and

King 1989, Swift 1984a), but reductions may be less for

poorer quality aggregates (Foltz and Truebe 1995).

In a study attempting to isolate the specific sources of

sediment, Burroughs and King (1989) identified the

cutslope, the roadway, and the fillslope (fig. 9.1). For each

of these components they suggested mitigation measures,

including application of mulch, geotextiles, seed, and sod.

Many other studies have demonstrated the effectiveness of

these treatments (table 9.2), and they are recommended in

many States. Luce and Black (1999), however, were not

able to measure any differences in sediment from roads for

bare and vegetated cutslopes of different heights in the

Oregon Coast Range. They concluded that the roadway and

the road ditch were the only significant sources of sediment.

Wemple and others (1996) and Elliot and Tysdal (1999)

found that the roads can influence a wider zone of erosion

than previously thought. Slopes and channels downhill from

the road can be sites of deposition, or the major source of

sediment from a given segment of road. The excess runoff

from roads can overload ephemeral channels, resulting in

significant downcutting of the channel.

Poor road drainage can also lead to saturation of road beds

and mass failure. In steep terrain, abandoned roads that do

not shed surface water can become saturated, increasing the

likelihood of failure. In areas of high rainfall, such as the

Coast Range in Washington and Oregon, more sediment

comes from roads due to landslides associated with roads

than from road surface erosion. Beschta (1978) reported that

watershed sediment yields increased from around 300 tons

per square mile [105 Mg per square kilometer (km)] per

year before roads and harvesting, to about 400 tons per

square mile (140 Mg per square km) per year after installing

roads and harvesting timber. Much of the increase in

watershed sediment yield in this high-rainfall area was from

mass failure. In a recent study in the Clearwater National

Forest in Idaho, 58 percent of the landslides that occurred

were associated with roads (McClelland and others 1998).

Recent studies in Oregon, however, suggest that road

impacts may have been overestimated (Robinson and others

1999), and that sediment from landslides in undisturbed

areas is similar to that in areas with roads. While surface

erosion is a chronic source of sediment associated with

numerous precipitation or snowmelt events every year,

landslides tend to contribute large amounts of sediment

during very wet years and no sediment during normal and

dry years. Landslide scars can also be sources of sediment

until they are revegetated. McClelland and others (1998)

calculated that the amount of sediment from the worst

landslides in 20 years was about 10 times a background

erosion rate, while the ongoing contribution from roads in

the basin was about 2.5 times the background rate.

In addition to roads, other rights-of-way such as pipelines,

are potential sources of sediment (Gray and Garcia-Lopez

1994, Sonett 1999). Any construction that exposes bare

mineral soil, particularly on sites that are adjacent to ditches

or streams, is likely to increase sedimentation. Once

Chapter 9
88

Roads and Other Corridors

Table 9.1—Typical erosion rates observed for different types of land use in the United States

Location Surface cover Erosion rate Reference

Tons/ac/yr

Eastern

watersheds Forests 0.003– 0.32 Patric 1976

Fernow NF, Observed bare and

West Virginia graveled roads 6.0 – 52.5 Kochenderfer and Helvey 1987

Appalachian Trail Trail 4 – 60 Burde and Renfro 1986

Southeast Roads 5 –144 Swift 1984a, 1984b

Southern Forests Trace – .32 Yoho 1980

watersheds Meadow .06 – .1

Prescribed burn .01 – .23

Careless clearcut 1.35

Roads 16 –150

Central Arkansas Roads 6.8 – 33.7 Beasley and others 1984

4 – 38.5 Miller and others 1985

Southeastern

Oklahoma Roads 8 – 77 Vowell 1985

Western

watersheds Rangeland .1 – 1.8 U.S. Department of Agriculture 1989

Northern Rockies Forests .04 Megahan 1974, McClelland and

others 1998

Forested watershed

 Undisturbed 0 Megahan and Kidd 1972b

 With roads .02

Roads 7.5 – 22 Ketcheson and Megahan 1996,

Megahan and Seyedbagheri 1986,

Megahan and Kidd 1972a

Washington

Olympics Roads 46 –550 Reid and Dunne 1984

Oregon Cascades Forested watershed .11 Fredrikson 1970 (most of road and

Roads added .56 harvest erosion attributed to

Harvested 18.4 landslides)

Roads .22 – 24 Foltz 1996

Oregon Coast

Range Roads 1 – 18 Luce and Black 1999

Oregon coast Forests .4 Beschta 1978

Northern Undisturbed forest .008 Rice and others 1979

California Coast After roads .63

Range watershed After roads and logging 1.9
89

Figure 9.2—Mean soil loss rates for four road surfaces before, during, and for 2 years after logging (based on Swift 1984a).

Chapter 9

installed, rights-of-way may continue to be sources of

sediment if revegetation or other erosion control practices

are not initiated (Gray and Garcia-Lopez 1994). Frequently,

off-road vehicle enthusiasts may use rights-of-way for

recreation. Also, mechanical and chemical control of

vegetation may reduce vegetative cover. Depending on the

site conditions, erosion rates from the compacted trails or

exposed rights-of-way may be similar to those of roads.

Trails for bicycling, walking, or horseback riding erode at

rates similar to roads (Leung and Marion 1996). The total

sediment delivered from these trails is generally lower,

however, because the total surface area of a narrow trail is

less than that of most roads.

Much of the sediment eroded from a right-of-way is rapidly

deposited below the right-of-way and never reaches a

stream. Rummer and others (1997) found no significant

sedimentation effects beyond the clearing limit of the road

in a bottomland hardwood study on a floodplain. Numerous

scientists have developed equations from field observations

to predict how far sediment will travel (Ketcheson and

Megahan 1996, McNulty and others 1995, Packer and

Christensen 1977, Swift 1986).

Various mitigation measures to reduce road erosion are

commonly prescribed by Federal and State agencies. The

most common methods include surfacing the road with

gravel, decreasing the spacing of cross drainage, locating

roads farther from streams, or limiting road gradients

(Burroughs and King 1989, Swift 1984a, Yoho 1980).

Table 9.2—Effectiveness of erosion mitigation techniques

Condition Reduction Reference

Percent

Erosion mat 74 – 99 Grace and others 1998

Seeding 82 – 95 Grace and others 1998

Grass on fillslope 46 – 81 Appelboom and others 1998

Straw and asphalt tack or 60 – 100 Burroughs and King 1989

erosion mats (depends

on percent cover)

Straw 60 – 80 Burroughs and King 1989

0

10

20

30

40

50

60

70

80

90

100

Prelogging Logging 1 year after

logging

2 years after

logging

Erosion rate (ton/acre)

Soil

Grass

6-in. gravel

8-in. rock
90

Roads and Other Corridors

Treatment of cut and fillslopes has also been effective in

reducing sediment delivery from new roads (table 9.2)

(Burroughs and King 1989, Grace and others 1998).

Sediment production can be reduced by applying higher

quality gravel (Foltz and Truebe 1995) or by reducing the

pressure in vehicle tires on the road network (Foltz 1994).

The installation of vegetated filter strips or slash filter

windrows below fills, or sediment basins below culverts, are

also management options that have reduced sedimentation.

In climates with distinct wet seasons, seasonal closure of

roads may be a desirable option to prevent rutting and

severe erosion.

In an effort to reduce the impacts of roads and railroads in

watersheds, many government management agencies are

removing unwanted corridors. In many national forests,

watershed restoration is synonymous with removal of excess

roads (Elliot and others 1996). Moll (1996) presented an

overview of road closure and obliteration methods in the

Forest Service. He recommended that watershed managers

consider access, drainage, erosion risk, slope stability, and

revegetation when planning any road closure or obliteration.

Table 9.3 summarizes management options for decommissioned roads. Elliot and others (1996) warn that the disturbances associated with road closure may cause more erosion

than simply abandoning a road that has been revegetated

and is hydrologically stable.

Surface erosion rates can drop significantly when roads are

closed. Figure 9.2 shows the relative impacts of different

road surfaces during the first 2 years after abandonment,

compared to erosion rates during construction and logging

(Swift 1984b). In the Oregon Cascades, Foltz (1996)

observed that during the first year of closure erosion rates

dropped from 4 to 0.5 tons per acre (9 to 1 Mg per hectare)

when marginal quality aggregate was applied, 20 to 2.5 tons

per acre (45 to 5.6 Mg per hectare) when good-quality

aggregate was applied. Erosion will often drop to background levels as the density of vegetation on an abandoned

road surface increases (Foltz 1996, Swift 1984b). Such a

decline is unlikely, however, if the abandoned road has

unvegetated surfaces and continues to concentrate runoff

water.

Several general principles can be applied to analyzing and

mitigating potential sediment sources from abandoned

corridors. The surface should be covered with vegetation. In

order to establish vegetation, it may be necessary to rip or

till the surface. In extreme cases, it may be necessary to add

topsoil. To encourage infiltration and revegetation, it may

be necessary to discourage off-road vehicle traffic by

installing permanent barriers to prevent wheeled access to

the corridor.

On abandoned roads, culverts can fail or become blocked,

causing ponding of water, embankment failure, and major

offsite sedimentation (Elliot and others 1994). Many older

roads or railroads were built with underdesigned culverts.

Some culverts were made from wood that is decaying or

metal that is corroding. In either case, most of these culverts

will eventually fail unless they are removed or replaced.

Culverts that are not regularly inspected can also become

blocked with woody debris or sediment. One of the most

common practices to minimize the risk of fill failure on

abandoned rights-of-way is to remove the culverts.

A number of prediction models have been developed to

estimate the amount of sediment that leaves forest roads.

Site-specific models were developed in the northern Rocky

Mountains by Forest Service hydrologists; the most recent is

the Watershed and Sediment Yield Model (WATSED) model

(U.S. Department of Agriculture, Forest Service 1990).

McNulty and others (1995) presented a Geographical

Information System-based method for predicting sediment

delivery from a road network, but they observed that

additional work with physically based models is necessary

to improve the prediction of sediment delivery from roads.

The physically based Water Erosion Prediction Project

(WEPP) model is under development for a wide range of

conditions including agriculture, range, and forest conditions (Laflen and others 1997). Because it is physically

based, the model can be applied whenever the factors that

cause erosion can be adequately described. Elliot and Hall

(1997) have developed a set of input templates for forest

roads and other disturbances. Elliot and others (1999b)

developed simplified tools based on the WEPP model to aid

managers in estimating the impacts of climate, soil, and

topography on the delivery of sediment from roads. These

models are available on the Web at http://forest.moscowfsl.

wsu.edu/fswepp/.

Reliability and Limitations of Findings

Researchers worldwide have measured increased sedimentation from roads and similar disturbances. The magnitude of

erosion varies considerably with climate, but the relative

impacts of soil, topography, and management are generally

the same (Elliot and others 1999b). Observed erosion rates

are highly variable (table 9.1) due to the high natural

variability in the factors that cause erosion. Even a welldesigned erosion experiment frequently results in variations

from the mean of up to 50 percent. This high variability

should be considered when interpreting any research or

monitoring results, or any erosion prediction value. Managers should exercise caution when applying any model to an

area where it has not received some validation. Predictive

technology for one climate, soil, and topography does not
91

Chapter 9

translate well to other conditions unless the model is able to

incorporate those site-specific characteristics.

The technology to remove abandoned roads is well established (Moll 1996). Numerous agencies including the Forest

Service (Moll 1996; U.S. Department of Agriculture, Forest

Service 1998) and the National Park Service2 have specialists to provide technical assistance in road closure, stabilization, revegetation, and removal. Many abandoned roads and

railroads require site-specific prescriptions for reclamation.

The same level of design that went into creating some of

these roads may be required to remove them (Elliot and

others 1996). Although this design expertise is available, the

cost may be prohibitive.

Research Needs

1. Upland erosion and sedimentation are well understood.

The long-term impacts of trapping sediment on hillsides

between sites of disturbance and streams and movement

of sediment within and through stream networks are not

well understood. Future research on overland transport

and storage of sediment and transport and storage in

stream networks will enhance sedimentation prediction.

2. Reports related to problems associated with abandoned

roads and railroads focus specifically on culverts or mass

failures. Surface and channel erosion may be a chronic

source of sediment for many years. Thus, published

information is frequently limited to episodic problems

rather than solutions to chronic problems. There is a need

for research to determine the probability of a failure

occurring as well as the probability that no failure will

occur. 2

 Spreiter, T.A. 1999. Road closure experiences of the National Park

Service. Presented at the Oregon and Washington engineering workshop;

1999 March 9; Gleneden Beach, OR. Oral presentation.

Table 9.3—Management options for decommissioned roads

Option Comment

Close road with barriers, vegetation,

ditch, or removal of first segment Recreational users may still obtain access.

Rip road surface Runoff is reduced (Luce 1997). Instability may be

increased (Elliot and others 1996).

Revegetate road surface See table 9.2

Remove culverts and restore channels Mitigation may be necessary on bare, excavated

embankments or in channels; remaining road

segments may not be accessible for future

maintenance (Moll 1996).

Reshape road surface to be outsloping

or partially recontoured with regular

waterbars Moll 1996

Install rock buttresses to stabilize cut

and fillslopes Moll 1996

Remove, recontour, or obliterate road Expensive [$0.60 to 1.50 per lineal ft (Moll

prism 1996)] Increase revegetation rate by excavating

until the old topsoil is reacheda

Mitigate obliterated road prism with

slash, mulch, geotextile, or seeding Moll 1996

a Spreiter, T.A. 1999. Road closure experiences of the National Park Service. Presented at the Oregon and Washington

engineering workshop; 1999 March 9; Gleneden Beach, OR. Oral presentation.
92

Roads and Other Corridors

3. Research is also needed to determine risks of failure or

erosion for specific road networks.

4. There is a need to develop field techniques to assist road

and watershed managers to make better decisions on

which segments of a road network are at the greatest risk

of a failure that may impact off-site water quality as well

as other resources.

5. Another need is to develop tools to estimate the amount

of sediment that may come from road closure activities,

both from reshaping or removing the road prism, and

from removal of stream-crossing structures.

Key Points

1. Roads and similar corridors can be a major source of

sediment in a forested watershed.

2. Effective measures to reduce sedimentation include

surface gravel, careful design of roads and water crossings, and removal of unwanted roads.

3. Abandoned roads may be sources of sediment if they

collect or divert surface runoff.

Hydrocarbons, Cations, and

Related Pollutants

Issues and Risks

Runoff from roads and similarly surfaced sites can contain a

host of hydrocarbons and other chemical pollutants,

adsorbed to sediments, as particles, or dissolved by the

runoff. These chemicals can find their way into surface and

subsurface water. Pesticides used to control unwanted weeds

can also be a source of pollution, and the reader should refer

to chapter 12 for further discussion.

Findings from Studies

Researchers have identified a range of chemicals in road

runoff (tables 9.4, 9.5). Some of the pollutants are from the

road material itself, some occur in the soil and rock on the

site and are released during construction or subsequent

erosion, and many are from vehicles. Traffic and road

surfacing may contribute undesirable cations, hydrocarbons,

and metals to surface and subsurface water (Maltby and

others 1995, Mungur and others 1994). Most studies on the

impact of roads and similar disturbances have focused on

heavily traveled roads such as major freeways (Mungur and

others 1994). If water source areas contain major roads,

runoff treatment may be necessary to ensure that undesirable

hydrocarbons do not enter the water supply.

Cations released from a road may have a buffering effect on

the runoff acidity, which may be beneficial in acid rain

areas. Morrison and others (1995) measured pH values from

6.0 to 7.0 in road runoff from small storms, compared to the

average rainfall pH of 4.1.

Ions from deicing or dust control chemicals are common

pollutants from road surfaces (Church and Friesz 1993,

Pugh and others 1996). Road salt contamination of shallow

ground water has become a serious problem, particularly in

the Northeast and Midwest (Church and Friesz 1993).

Church and Friesz (1993) state that during a 7-year period in

Massachusetts, there were complaints from 100 of the 341

municipalities about road salt contamination. Nationally,

about $10 million are spent each year for prevention or

remediation of problems associated with road salt contamination. Surface water is less vulnerable to such contamination than ground water, because there tends to be much

greater dilution and mixing in turbulent channels carrying

runoff from roads (Jongedyk and Bank 1999). Calcium

magnesium acetate and potassium acetate are deicing

chemicals with less serious environmental consequences

than sodium chloride because they contain weak biodegradable acids. Sodium chloride, calcium chloride, and magnesium chloride, however, leave residues of chloride ions that

may contaminate ground water (Jongedyk and Bank 1999).

Some of these ions (calcium, magnesium, and potassium)

can enhance vegetation growth along highways (Pugh and

others 1996). In some cases, elevated levels of deicing

cations such as sodium in the road runoff, may be adsorbed

by the soils near the road, and pose no further concern to the

aquatic ecosystem (Shanley 1994). Pugh and others (1996)

observed that ion concentrations from an adjacent interstate

highway decreased exponentially with distance from the

road in a peat bog. Although many thousands of tons of salt

are spread annually on highways, because of dilution, salts

in runoff are not likely to be a major source of pollution for

drinking water except where they use shallow ground water

even though impacts on the aquatic ecosystem may be great.

Road dust can transport unwanted chemicals to surface

water. Christensen and others (1997) observed recent

accumulations of polycyclic aromatic hydrocarbons (PAH’s)

in a Wisconsin stream, and identified dust from nearby roads

as the source of the pollutant.

Oil-based dust suppressants may be environmentally more

risky than salt-based products. A literature search for the

Forest Service (Heffner 1997) found reports that calcium

and magnesium chloride showed some toxicity towards
93

Table 9.5—Mean concentrations of a number of

pollutants in highway runoff in Minnesota

Pollutant Range Mean

- - - - - Milligrams per liter - - - - - -

Total nitrogen 0.6 – 8.14 1.67

Chloride 1 – 46 000 1 802

Sulphate 5 – 650 45

Sodium 2 – 67 000 3 033

Total phosphorous .06 – 7.8 .6

BOD 1 – 60 12.6

COD 2 – 3 380 207

Total suspended solids 8 – 950 118

Total dissolved solids 22 – 81 700 10 440

- - - - - Micrograms per liter - - - - - -

Chromium 1.5 – 110 13

Copper 3 – 780 47

Iron 180 – 45 000 4 162

Lead 11 – 2 100 207

Zinc 10 – 1 200 174

Nickel 1 – 57 10

Cadmium .2 – 12 1.7

Mercury .08 – 5.6 .49

Aluminum 30 – 14 000 2 694

Arsenic .1 – 340 19

BOD = biological oxygen demand; COD = chemical oxygen demand.

Source: Thomson and others 1997.

Chapter 9

plants, whereas ligninsulfonate increased water biological

oxygen demand. The study concluded:

based on the literature review and typical application

rates for dust abatement, the effects of these compounds on plants and animals would be negligible.

For the purposes that the Forest Service uses these

compounds, the selection of one over another would

be more dependent on cost, availability, and local

conditions than effects to the environment.

Some dust inhibitors may also decrease road erosion,

decreasing the likelihood of off-site transport of sediment

and related pollutants (Ice 1982).

Chemicals used to preserve utility poles and railway

crossties are potential sources of pollution. Wan (1994)

found that concentrations of PAH’s in the soil were higher in

the immediate vicinity of utility poles than on surrounding

farm or rangeland. Soil concentrations of PAH’s dropped

rapidly from 550 micrograms (μg) per liter to 23.2 μg per

liter within 13 feet (4 meters) of a treated pole. Background

levels were between zero and 0.8 μg per liter. Such findings

emphasize the importance of maintaining vegetated buffers

to reduce transport by erosion of contaminated soil between

rights-of-way and any sensitive water resources.

Measuring concentrations of many pollutants is tedious and

expensive. To reduce the cost, surrogate relationships have

been developed between more easily measured pollutants,

such as suspended solids (mainly sediment) or dissolved

Table 9.4—Pollutants that have been observed in runoff from road surfaces

Pollutant Comment Reference

Cd, Cu, Pb, Zn Treated in wetlands Mungur and others 1994

Highway deicing salt Na adsorbed in soil Shanley 1994

Cu, Zn, hydrocarbons, PAH’s Accumulated in aquatic biota Maltby and others 1995

PAH’s Altered aquatic communities Boxall and Maltby 1997

Ca, Mg, Na, K, Cl Captured in peat bogs Pugh and others 1996

Total petroleum

hydrocarbons, Pb, Zn Reduced by vegetation Ellis and others 1994

Heavy metals, petroleum hydro- Treatment ponds can remove Karouna-Renier and Sparling 1997

carbons, pesticides, sediment, up to 95 percent of pollutants.

nutrients

Ca = calcium; Cd = cadmium; Cl = chloride; Cu = copper; K = potassium; Mg = magnesium; Na = sodium; PAH’s = polycyclic aromatic

hydrocarbons; Pb = lead; Zn = zinc.
94

Roads and Other Corridors

solids and other pollutants that are difficult to measure

(Thomson and others 1997). Such surrogates may be useful

if relationships were developed for nearby conditions, but

they become less reliable when extrapolated to other

regions.

Gilson and others (1994) completed research on the effectiveness of filter systems for highway runoff to improve

surface water quality in the karst terrain in Texas. They

found that some alternatives to sand filters have higher

adsorptive capacities initially, but filtration efficiencies

tended to approach that of sand after several runoff events.

A Virginia study found that highways in karst areas should

be located to avoid polluting surface water that drains into

caves (Hubbard and Balfour 1993). This study found raw

sewage and petroleum fumes in the cave system. Keith

(1996) described extra precautions on road location and

drainage designs that were taken in Indiana to minimize the

ground water impact of a new road design in a karst area.

Pollutants in runoff can be trapped in natural or artificial

wetland areas (Ellis and others 1994, Karouna-Renier and

Sparling 1997, Mungur and others 1994). Karouna-Renier

and Sparling (1997) found that such treatment systems could

remove up to 95 percent of metals, nutrients, and sediment.

Monitoring of the performance of such areas is necessary to

ensure that they are functioning as desired (Startin and

Lansdown 1994).

Another treatment method is a partial exfiltration trench.

This type of device filters out the suspended solids that carry

many of the undesirable metals and hydrocarbons from road

surfaces (Sansalone and Buchberger 1995). The trench

improved the quality of both rainfall and snowmelt runoff

from roads. Because of the wide range of runoff rates,

however, multiple treatment methods may be necessary to

decrease the pollutant load from large as well as small

storms (Romero-Lozano 1995). Detention basins are needed

to catch the first flush of highly polluted runoff. A filtration

system is needed to treat the runoff from later in the storm,

which is likely to be at a higher flow rate, but requires less

treatment.

Sediment basins and similar structures built to contain

polluted road runoff can become sources of pollution

through seepage into the ground, or through other forms of

hydraulic or structural failure. In either case, sediments with

large amounts of adsorbed chemicals can enter a stream.

The pollutants can become concentrated in these basins,

increasing the risk of offsite pollution (Grasso and others

1997, Morrison and others 1995). Grasso and others (1997)

observed a lead content of 1392 milligrams per kilogram on

one site and recommended soil washing be carried out to

prevent offsite pollution. One of the best defenses against

such risks is cleaning and maintaining such structures to

minimize the risk of failure.

Past designs of runoff structures tended to collect water and

route it directly to a stream. New designs that disperse water

to ensure greater infiltration and onsite attenuation of

pollutants can improve runoff quality (Elliot and Tysdal

1999, Li and others 1998). Not all sites lend themselves to

this approach, particularly where rights-of-way are limited.

Another recent innovation to reduce offsite pollution from

roads and similar areas is to surface them with permeable

pavement (Church and Friesz 1993). Permeable pavement

combined with high-infiltration shoulders significantly

decreased salt content in nearby ground water (Church and

Friesz 1993, Jongedyk and Bank 1999). European researchers found that permeable pavement significantly reduced

outflow levels of lead and suspended solids.

Reliability and Limitations of Findings

Much of the research associated with chemical pollution

from roads has taken place near large urban centers. The

findings are generally reliable for their locality, but care

needs to be taken in extrapolating to other conditions,

particularly nonurban areas. The water-quality risks associated with hydrocarbon pollution are closely linked to the

density of traffic. Watersheds with minimal traffic are

unlikely to experience any of the problems discussed. These

results should only be applied to more remote watersheds

with caution and some form of monitoring.

Research Need

Pollution from main roads that cross sensitive forest and

grassland watersheds should be measured. Quantitative data

are needed on the benefits of dust abatement chemicals for

reducing erosion and pollution of streams near roads.

Key Point

Many pollutants from vehicles, deicing and dust abatement

chemicals, and road surfacing material have been measured

in runoff from roads. Most of these measurements have been

from roads with heavy traffic. Some level of monitoring

may be necessary to determine pollution problems. Levels

of pollution can often be related to levels of easily measured

sediment concentration. Some cations in runoff may be

beneficial in buffering acid rain. There are methods to

collect and treat or to harmlessly disperse polluted road

runoff.
95

Chapter 9

Fuels and Other Contaminants

from Accidental Spills

Issues and Risks

Accidents are rare on low-use roads and other rights-of-way

in remote watersheds. Risks of an accident causing contamination spills are related to the traffic density, quality of road,

and frequency of contaminant transport. Railroads pose

similar risks, particularly on aging lines, or on busy routes

linking industrial centers.

Findings from Studies

Hazardous chemical spills from vehicle accidents can pose a

direct, acute threat of contamination to streams. Risk

analysis models have been developed for busy paved roads

in nonmountainous terrain, but these models are seldom

applicable to low traffic, remote watersheds. Chemicals that

may be spilled include fuel, fertilizer, pesticides, and mining

chemicals (U.S. Department of Agriculture, Forest Service

1998). Airfields can often be sources of ground water

contamination due to spills of fuels and other material

(Levine and others 1997).

Accidents may occur anywhere along a given road or

railroad, but stream crossings and bridges tend to be

frequent sites of accidents due to damage by floods, or a

narrowing of the roadway. Whether the pollutant is able to

reach a nearby stream is an important concern. Spills at

stream crossings have a high likelihood of reaching surface

water because of its close proximity. Frequently, transport of

the pollutant overland, or through the soil, depends on the

local climate, season, and hydrology.

Reliability and Limitations of Findings

There is little information available on the risk of accidental

spills in remote areas. Whatever information can be found is

likely to be site-specific, and judgment must be used to

apply it elsewhere. Watershed managers will need to

develop their own set of potential risks, based on local

conditions. Along with those risks, they will need to develop

a set of potential mitigation measures, both in the water

source area, and in the treatment system.

Ability to Address Issues

Most counties have established committees to address local

emergencies or disasters. An accident that impacts a local

water supply is a prime example of such an emergency.

Water supply managers should work with local emergency

or disaster committees or services to ensure that mitigation

plans and equipment are in place to deal with toxic spills

that may occur near a water source.

Research Need

Because of the site-specific nature of this risk, it is difficult

to define a broad research activity for remote watersheds. It

is likely that research will continue to study risks associated

with busier roads, so monitoring of those results for application to remote watersheds may be beneficial.

Key Point

The risk of vehicle accidents and spills depends on road

hazards and traffic volume. Watershed managers need to

evaluate risks on a given watershed and develop prevention

or mitigation measures specific to their own conditions.

Pipeline Failures

Issues and Risks

Pipelines carrying a wide range of substances, including

drinking water, sewage, and petroleum products, can fail,

leading to pollution of both surface water and ground water.

In the past 15 years, about 200 oil pipeline failures have

occurred per year, with an average net loss of about 600

barrels (95 cubic meters) for each spill (U.S. Department of

Transportation 1999).

Findings from Studies

Pipelines tend to have fewer accidents and injuries than

other modes of transport (Jones and Wishart 1996). To

minimize pollution impacts, most modern pipeline systems

are equipped with devices to quickly shut down pumping if

a change in flow or pressure is sensed (Ariman 1990).

A number of disturbances increase the likelihood of pipeline

accidents (fig. 9.3). Road or construction accidents and

damage from boulders are common external causes of

damaged pipelines (Driver and Zimmerman 1998, Stalder

1997). Areas prone to severe erosion, landslides, and

earthquakes tend to have more accidents (Ariman 1990,

Gray and Garcia-Lopez 1994, Hart and others 1995). For

example, Hart and others (1995) predicted that the probability of rupture for a pipeline in California increased from 0.0

for earthquakes with a magnitude below 5 to 1.0 for

earthquake magnitudes greater than 6.0. They also predicted

other probabilities of failure based on pipe length and

installation. They recommended numerous design measures

including depth of burial, trench design, and pipe wall

thickness, to minimize failure due to earthquakes.
96

Corrosion

24%

Other

30%

Equipment

6%

Damage

26%

Pipe

4%

Operation

5%

Welds

5%

Figure 9.3—Summary of causes of pipeline accidents in 1998 in the United

States (U.S. Department of Transportation 1999).

Roads and Other Corridors

Pipelines sometimes fail at river crossings due to erosion of

the streambank or bottom (Doeing and others 1995, Teal and

others 1995). Pipelines carrying sewage and industrial

wastes are frequently located in floodplains and are at

particular risk from flood damage, or from overloading due

to high runoff rates. Disturbances in a watershed, such as a

fire, may cause landslides that lead to pipeline failure.

Failure of water supply pipelines or canals can lead to

considerable erosion if controls to monitor flow conditions

are not in place.

Soil shrinking and swelling and freezing and thawing can

lead to pipeline fatigue and premature failure. Corrosion due

to electrolysis (Stalder 1997) and stress corrosion cracking

can also occur on older pipelines (Wilson 1996). Aboveground pipelines can fail due to wind fatigue (Honegger and

others 1985). Any pipeline may experience seam failure

(Yaorong and others 1996).

Risk assessment models to aid in pipeline design and

operation have been developed (Hart and others 1995,

Nessim and Stephens 1998). These models can identify

segments of pipe most at risk of failure, and maintenance

can be concentrated on those segments. Risk rates of 0.0022

spills per mile per year are quoted in one environmental

assessment (U.S. Department of the Interior, Bureau of Land

Management 1978).

Table 9.6 shows the extent of contamination from 53 oil

pipeline spills. The extent depends on the pipeline characteristics and on the soils and terrain. Risks of failure from

normal, predictable events can be reduced to almost zero

with adequate pipeline monitoring (Stalder 1997). In

addition, technologies have been developed to mitigate the

impacts of spills quickly and effectively (Sittig 1978).

Pipeline failures can pollute ground water as well as surface

water. Petroleum products tend to float on ground water, but

the processes associated with breakdown of oil underground

are not well understood. Underground methane generation

by anaerobic bacteria is common after a pipeline break.

Substantial amounts of the volatile petroleum hydrocarbons are transported from the surface of the water

table through the unsaturated zone as vapor, which

subsequently dissipates to the atmosphere or is

biodegaded (Revesz and others 1996).

Eganhouse and others (1996) observed that an underground

breakdown process from microbial degradation leads to the

detection of a plume containing aliphatic, aromatic, and

alicyclic hydrocarbons.

Pipelines carrying water and sewage may also be present on

watersheds. Although the same principles of failure apply to

these pipelines, they are generally not as well monitored and

may be older and more prone to failure.

Reliability and Limitations of Findings

These findings are generally reliable because much of the

pipeline industry receives close government scrutiny.

Pipeline failures tend to be mechanical and predictable and

findings are generally applicable to local conditions.

Current technology can address the risks associated with

pipeline failure. Technologies to minimize pipeline failure

are well established in the petroleum industry, as are

controls to minimize pollution of surface and ground water

should a failure occur. These same technologies can also be

applied to other pipelines in sensitive watersheds. Managers

of watersheds containing pipelines should work with

Table 9.6—Extent of soil contamination by 53 oil spills of

various sizes in Alberta, Canada

Average Average Average film

volume area thickness

Barrels Ft2 In.

54 8,000 0.4

880 70,000 .8

13,200 600,000 1.6

Barrel = 42 gallons of petroleum.

Source: U.S. Department of the Interior, Bureau of Land Management 1978.
97

Chapter 9

pipeline managers to minimize risks to water supplies. In

addition, the U.S. Department of Transportation has an

Office of Pipeline Safety (OPS) to assist in addressing risks

associated with pipelines. One of its responsibilities is to

identify areas that are unusually sensitive to a hazardous

liquid pipeline release. The OPS has an ongoing program

that may assist watershed managers in risk management (see

Web site).

Research Need

The oil industry has developed sophisticated systems for

managing pipelines. There is a need to develop similar, but

less costly, technologies for water and sewer pipelines in

sensitive watersheds. The fate of oil pollution in the ground

is not well understood, and further research is needed to

better understand the chemical and biological processes

associated with degradation of petroleum products on and in

the soil.

Key Point

Causes of failures on petroleum pipelines are well understood, and controls are generally in place to minimize

environmental risks of failures. Such measures are less

developed for water and sewage pipelines, so the risks of the

failure of such systems may be higher.

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mountain roads. Forest Science. 30(3): 657–670.

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the Southern Appalachian Mountains. Southern Journal of Applied

Forestry. 8(4): 209–215.

Swift, L.W., Jr. 1986. Filter strip widths for forest roads in the Southern

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U.S. Department of Agriculture. 1989. The 2d RCA appraisal. Soil, water,

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U.S. Department of Agriculture, Forest Service. 1998. Forest Service

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Roads and Other Corridors
101

Part III:

Effects of Vegetation

Management on

Water Quality

Logs are kept wet at the woodyard, Mississippi. Photo by Bill Lea
103

Chapter 10

Timber Management

John D. Stednick1

Introduction

Forest management activities that disturb the soil or remove

vegetation may potentially affect the quality of drinking

water sources. Examples include removing trees from the

site for timber harvest, forest stand regeneration, and stand

improvement. Soil disturbance from tree felling is minor,

but movement of logs or whole trees to a landing or collection point may disturb the soil surface. Other soil surface

disturbances may be related to collection and haul roads.

Roads are addressed separately in chapter 9. Stand improvement may include selective harvesting of trees in either

dominant or suppressed crown positions. Forest stand

thinning may increase water and nutrient availability, but

any increase is utilized quickly by the remaining vegetation.

Stand improvement may also include subordinate vegetation

removal by fire (see chapter 12) or by herbicides (see

chapter 13).

This chapter reviews the potential effects of timber management on water quality. Forest vegetation management may

affect concentrations of suspended sediment and nutrients in

surface water and stream temperature.

Erosion/Sedimentation

Forest management activities associated with timber

harvesting may affect the physical, chemical, and biological

properties of the soil. If these activities increase soil erosion,

then water quality may be decreased through suspended

sediment transport or stream sedimentation. Soil erosion is

the detachment and movement of soil particles. It is measured as tons per acre per year [metric tonnes (Mg) per

hectare per year]. Suspended sediment is eroded soil

material transported in the water column of a stream. It is

measured as a concentration such as milligrams per liter or

as turbidity, which is an optical measurement of the water’s

ability to diffract light and is expressed as nephelometric

turbidity units (Stednick 1991).

Site properties that affect erosion processes include vegetative cover, soil texture, soil moisture, and slope, among

others (Falletti 1977, Renfro 1975). The sediment load of

streams (both suspended and bed load) is determined by

such characteristics of the drainage basin as geology,

vegetation, precipitation, topography, and land use. Sediment enters the stream system through erosion processes.

To achieve stream stability, an equilibrium must be sustained between sediment entering the stream and sediment

transported through the channel. A land-use activity that

significantly changes sediment load can upset this balance

and result in physical and biological changes in the stream

system (State of Idaho 1987).

The existing form and characteristics of streams have

developed in a predictable manner as a result of the water

and sediment load from upstream. Natural channels are selfformed and self-maintained. Both water and sediment yields

may change due to timber management or other land-use

activities upstream.

Issues and Risks

The forest practices with the greatest potential for causing

erosion and stream sedimentation are road construction,

tractor skidding of logs, and intensive site preparation.

These activities can contribute to surface, gully, and largemass soil movements (see chapters 3, 9). Other soil erosion

processes may occur at smaller scales and rates. Generally,

as site disturbance increases, soil erosion increases.

Most soil erosion studies only measure the amount of soil

moved or displaced. The actual amount of eroded soil

reaching the surface water is a small percent (2 to 10

percent) of the erosion occurring in the watershed. This

percentage is termed the sediment delivery ratio and is the

amount of sediment produced divided by the amount of soil

erosion as a function of the watershed area (Dunne and

Leopold 1979). Soil erosion and subsequent sediment

delivery to the stream usually occurs at a specific location or

locations downstream from the disturbance.

Sediment accumulation in stream channels may adversely

affect water quality and aquatic life. Stream sedimentation 1 Professor and Program Leader, Colorado State University, Watershed

Science, Fort Collins, CO.
104

Timber Management

may adversely affect stream macroinvertebrates, intergravel

dissolved oxygen, and intergravel flow and migration paths.

When waters with increased sediment or turbidity are used

for drinking water, treatment costs increase. The water must

be filtered or stored to allow settling to remove suspended

sediment. Often chlorination rates must be increased to

disinfect water with elevated suspended sediment because

bacteria may be associated with the sediment. See chapter 3

for further discussion of sediment effects.

Findings from Studies

Undisturbed forest watersheds usually have erosion rates

from near 0 to 0.25 tons per acre per year (0.57 Mg per

hectare per year) (Binkley and Brown 1993a). Erosion rates

have been estimated as <0.1 tons per acre per year (0.2 Mg

per hectare per year) for three-quarters of eastern and

interior western forests (Patric and others 1984). Typical

timber harvesting and road construction activities may

increase erosion rates to 0.05 to 0.25 tons per acre per year

(0.11 to 0.57 Mg per hectare per year) (table 10.1). More

intensive site preparation treatments such as slash windrowing, stump shearing, or roller chopping may increase soil

erosion rates by up to 5 tons per acre per year (11.4 Mg per

hectare per year). Erosion from unpaved road and trail

surfaces may be higher yet (see chapter 9).

Numerous studies have been done on the effects of different

forest management practices on erosion rates or sediment

production (table 10.1). In general, increased site disturbance will result in increased soil erosion and subsequent

sediment production. The type and magnitude of erosion

depend on the amount of soil exposed by management

practices, the kind of soil, steepness of the slope, weather

conditions, and any treatments after the disturbance (Swank

and others 1989).

Logging in the Southeastern United States increased erosion

to 1.8 tons per acre per year (4.1 Mg per hectare per year)

from the undisturbed rate of 0.005 tons per acre per year

(0.011 Mg per hectare per year); about 10 percent of the

increase was attributed to site preparation (Hewlett 1979).

Roller chopping and slash burning in North Carolina had

little effect on soil erosion after harvest, but soil disking and

herbicide application increased soil erosion to 4.5 tons per

acre per year (10 Mg per hectare per year) (Pye and

Vitousek 1985).

Timber harvesting and subsequent yarding can increase

sediment in streams by increasing surface erosion rates and

increasing the risk of mass soil movement (Brown and

Krygier 1971, Brown and others 1976, Davis 1976). Site

disturbance can reduce infiltration rates and increase

overland runoff and related surface erosion.

Logs are moved (skidded) from the stump to a landing by

tractor, cable, aerial systems, or animals. Tractor skidders

may be either crawler or wheeled units, both of which are

frequently equipped with arches for reducing the extent of

contact between log and ground. Site disturbance will vary

greatly with the type of skidding or yarding system. Crawler

tractors generally cause the greatest amount of site disturbance, followed closely by wheeled skidders, but on some

sites use of wheeled skidders can result in more compaction

than crawler tractors (Bell and others 1974, Davis 1976).

One method of decreasing the amount of soil disturbed by

crawler tractors or wheeled skidders is through careful

layout of skid trails (Rothwell 1971). Careful location of

skidroads can greatly decrease the impact of tractor logging.

Cable logging systems will result in less site disturbance

because yarding trails are established to the yarding tower

machinery, which is restricted to road surfaces. Cable

systems can be ranked in order of decreasing soil disturbance as follows: single drum jammer, high lead cable,

skyline, and balloon (Brown and others 1976, Davis 1976,

Stone 1973). Helicopters and balloons will likely result in

minimum site disturbance, but both are costly and subject to

operational constraints.

Unlike many other land uses that disturb soil for long

periods, any increase in sediment yields from timber

management activities is usually short-lived. Surface soil

disturbances provide a sediment supply, but once the finer

materials are transported and as revegetation occurs, that

site is less apt to continue eroding. Sediment yields or

measured suspended sediment concentrations decrease over

time as a negative exponential (Beschta 1978, Leaf 1974,

Megahan 1975, NCASI 1999a). This time factor should be

considered when assessing watersheds for impacts on

drinking water (Stednick 1987). Swank discusses sediment

yields over time as the forest succession after logging

proceeds (see chapter 11).

Raindrop splash may potentially sort surface soil particles

and create an armor layer or erosion pavement. Erosion

pavements can form quickly on some soils in the West,

discouraging further erosion. In the South, however, many

surface soils have fine texture to depths of several inches

(centimeters) to several feet (meters). There, the soil surface

often becomes sealed, accelerating surface runoff, erosion,

and sedimentation. Fine soil particles continue to be

transported by surface runoff until the area is completely

revegetated. Revegetation may take 2 years where trees

have been harvested, 3 to 5 years for skid trails and temporary logging roads, and 3 to 5 years for site preparation

depending on the type of practice.
105

Table 10.1—Effects of various timber harvests or site preparations on soil erosion and sediment production

Erosion or

Location Treatment sedimentation rate Unit Measurement Reference

East

Georgia Piedmont Control 0.002 Tons/ac/yr Sediment Hewlett 1979

Harvest 1.8 Tons/ac/yr

Roading 1.6 Tons/ac/yr

North Carolina Roller-chop and burn 1.8 Tons/ac/yr Erosion traps Pye and Vitousek 1985

Shear stumps, windrow slash 1.8 Tons/ac/yr

Above plus herbicide (glyphosate) 4.5 Tons/ac

Southeast Natural 0 – .2 Tons/ac/yr Erosion Burger 1983

United States Harvest with roads .05 – .23 Tons/ac/yr

Burn .02 – .32 Tons/ac/yr

Chop .02 – .11 Tons/ac/yr

Chop and burn .07 – .18 Tons/ac/yr

Windrow slash .09 – .11 Tons/ac/yr

Disk 1.13 – 4.5 Tons/ac/yr

Coweeta Hydrologic Poor road design 840 Yd3/mi of road Volume/road length Swift 1988

Laboratory, NC Poor road design 5700 mg/L Sediment conc.

Good road design with grass or gravel .02 – .3 Tons/ac

Hubbard Brook Natural (WS4-WS6) .011 Tons/ac/yr Erosion Hornbeck and others 1987

Experimental Forest, Harvest and herbicide .05 Tons/ac/yr

NH

Fernow Experimental Natural .05 Tons/ac/yr Erosion to stream Aubertin and Patric 1972, 1974

Forest, WV Harvest .02 Tons/ac/yr

Cherokee County, Natural .008 Tons/ac/yr 4-yr average Blackburn and others 1986

TX Harvest, chop, and burn .006 Tons/ac/yr Blackburn and Wood 1990

Harvest, shear, windrow, and burn .36 Tons/ac/yr

Natchez, TN Control 82 mg/L Stormflow sediment McClurkin and others 1985

Harvest only 183 mg/L concentration

Gulf coastal Control 0.2 Tons/ac/2 yr Erosion over 2 yr Beasley 1979

Mississippi Harvest, chop, and burn 6.7 Tons/ac/2 yr

Harvest, shear, and windrow 6.7 Tons/ac/2 yr

Harvest, shear, and windrow 9.0 Tons/ac/2 yr

and plow beds

continued

Chapter 10
106

Timber Management Table 10.1—Effects of various timber harvests or site preparations on soil erosion and sediment production (continued) Erosion or Location Treatment sedimentation rate Unit Measurement Reference East (continued)

South Carolina Clearcut .07 Tons/ac/yr Erosion Van Lear and others 1985

Control .01 Tons/ac/yr

Sumter National Low-intensity burn .06 Tons/ac/yr Erosion Robichand and Waldrop 1994

Forest, SC High-intensity burn 2.6 Tons/ac/yr

West

Oregon Roads >.25 In./yr Road surface Fredriksen 1965

Fraser Experimental Control .018 Tons/ac/yr To stream Leaf 1974

Forest, CO Roads and harvest—Fool Creek .005 Tons/ac/yr Leaf 1974

Roads and harvest—Deadhorse .014 Tons/ac/yr Stottlemeyer 1987

H.J. Andrews Control WS9 .014 Tons/ac/yr To stream Sollins and others 1980

Experimental Clearcut WS10 .09 Tons/ac/yr

Forest, OR

Ouachita Control .005 Tons/ac/yr Surface erosion Miller and others 1988

Mountains, AR Harvest, roller-chop, and burn .09 Tons/ac/yr

Harvest, selection cut .018 Tons/ac/yr

Silver Creek, ID Clearcut with buffer 5.8 Tons/ac/yr—1st yr Erosion Clayton and Kennedy 1985

1.8 Tons/ac/yr—2d yr

Beaver Creek, Control .009 – .09 Tons/ac/yr Annual range Ward and Baker 1984

AZ Control 11 Tons/ac/yr Observed maximum

31% clearcut 1.3 Tons/ac/yr Observed maximum

100% clearcut 27.4 Tons/ac/yr Observed maximum

Clearcut .14 Tons/ac/yr Plot erosion Heede and King 1990

Caspar Creek, CA Selective clearcut 4.5 Yd3/ac/yr Erosion Krammes and Burns 1973

Drew County, AR Clearcut .120 Tons/ac/yr Erosion Beasley and Granillo 1988

Selection cut .005 Tons/ac/yr

Control .002 Tons/ac/yr

Clark County, AR Clearcut—mechanical prep .24 Tons/ac/yr Erosion Beasley and others 1986

Clearcut—chemical prep .11 Tons/ac/yr

Control .03 Tons/ac/yr

Conc. = concentration, WS = watershed.
107

Some form of site preparation is often needed to ensure the

establishment of tree reproduction after timber harvest. The

purpose of site preparation is to provide the environmental

conditions necessary for seed or seedling survival and early

growth. Site preparation usually involves providing a

mineral seedbed and controlling competing and nondesirable vegetation. Site preparation treatments include

fire, herbicide application, slashing and windrowing, roller

chopping, soil disking, or other mechanical techniques.

Fertilizer may be applied to help establish seedlings and to

speed their growth after establishment.

In the Southeastern United States, upland hardwood stands

are sometimes converted to pine (Pinus spp.). Site preparation treatments include burning or chemical treatments to

kill the existing vegetation. Soils in the region are often fine

textured and deep and may continue to erode at an accelerated rate for a few years. A winter burn and herbicide

application increased stormflows, overland flows, peakflows, and sediment production from two small watersheds

in northern Mississippi (Ursic 1970). Three years after the

fire, when monitoring ended, most of the hydrologic effects

were still evident.

Suspended sediment transport varies with the areal extent of

the soil disturbance, nearness of a stream, and stream

energy. Suspended sediments are often fine-textured

materials with large surface areas per unit of weight. These

large surface areas are reactive and may adsorb and absorb

various constituents including phosphorus, introduced

chemicals, and petroleum products.

Streamside vegetation or filter strips have been used to

prevent overland flow and soil erosion from reaching

surface waters. The filter strip, or equivalent, decreases the

velocity of the overland flow by surface roughness. The

decreased velocity allows sediment to settle out and overland waters to infiltrate into the undisturbed soils. The

streamside vegetation filters were originally used to control

or limit road-derived sediment from reaching forest streams.

The filter was a recommended width and was dependent on

hillslope. These filter strips are effective in sediment

removal unless an extreme precipitation or overland flow

event exceeds the sediment detention/retention capacity. The

characteristics that make filter strips work include width,

vegetative and litter cover, surface roughness, and microtopography. Microtopography allows overland flow to

concentrate in certain areas and flowpaths. Control of roadderived sediment migration is frequently by these strips. The

effectiveness of filter strips on controlling soil erosion for

most harvest and site preparation practices has not been

rigorously tested.

Routing and storage are particularly important components

in the transport of sediment through the stream system. They

are critical to the quantification of short- and long-term

impacts of land-use activities on the quality of drinking

water sources. However, the storage and routing processes

are highly variable and do not exhibit steady-state behavior

(see chapter 3).

Catchment studies have identified correlations between

annual peak discharge and annual sediment discharge and

between total annual flow and annual sediment discharge

(NCASI 1999a). Altering flow and erosion may upset

channel stability, increasing turbidity and sediment concentrations to drinking water sources.

Reliability and Limitations of Findings

Studies have shown that increased site disturbance has the

potential to increase soil erosion and sediment production.

Soil erosion and sediment yield from undisturbed forest

watersheds are low. Site disturbance from timber harvesting

activities vary by logging and yarding techniques, site

preparation practice, operator techniques, soil vegetative

cover, slope, soil moisture, soil depth, and soil texture

among other environmental factors. Soil erosion processes

are well understood, and models have been developed for

regional predictions of soil erosion throughout the United

States.

Measuring instantaneous sediment concentration (and

turbidity) in small streams is relatively easy. Measuring soil

erosion is not. Erosion is variable in time and space, and the

eroded soil must reach the stream channel to become

sediment. Once in the stream channel, most of the sediment

is transported irregularly when streamflows are high.

Sediments may be stored in the channel and released over a

long period. In-channel disturbances may create in-channel

sediment sources, separate from the hillslope processes.

Large sediment inputs to stream channels can be assessed by

monitoring the physical features of the channel (MacDonald

and others 1991, State of Idaho 1987). Such features include

channel width-to-depth ratios, pool volume occupied by

sediment, and substrate size and particle size distribution.

Research Needs

1. There is no standard or protocol for erosion plot research

on forest land. A standard research method for soil

erosion studies should be decided upon.

2. The importance of dry ravel as an agent of erosion needs

further investigation.

Chapter 10
108

Timber Management

3. Research is needed on routing eroded soil to streams.

Erosion does not equal suspended sediment. Measured

erosion rates do not or should not imply that eroded soil

is reaching the stream channel. Suspended sediment

monitoring is not difficult, but requires labor and equipment that may not be available. For source areas the

question is: Do suspended sediment concentrations best

measure the effects of site disturbance?

4. Recommendation and design of vegetative filter strips are

often based on width only. Research needs to better

define the characteristics that control sediment movement

including slope, vegetative and litter cover, runoff

velocity and volume, surface roughness, and microtopography of the filter strip and disturbed area above.

5. Research is needed on monitoring of stream channel

geomorphologic features, which may provide a good

measure of land-use effects, particularly multiple or

cumulative effects. Increased annual water yield from

timber harvesting has been well documented, but the

effect of timber harvesting on peak flows is less clear.

Can this altered hydrology increase sediment transport

from in-channel sources and result in changes in channel

morphology? Conversely, how much increased sediment

input can a stream segment receive without changes in

channel morphology?

Key Points

Site disturbance may result in soil compaction and decreased infiltration capacity. If infiltration capacity is

exceeded by precipitation intensity, overland flow may

result in soil erosion and suspended sediment production.

Even undisturbed forest watersheds produce sediment,

mostly from in-channel sources. Sediment impacts from

timber management activities can be minimized by:

1. Careful planning, supervising, and implementing of

forest practices.

2. Keeping the treatment area small and hydrologically

isolated.

3. Leaving adequate filter strips between treatment areas

and streams.

4. Maintaining ground cover in the treatment area to reduce

surface runoff and erosion, and increasing the effectiveness of filter strips to trap eroded soil before it enters the

stream.

5. Operating during the season with the lowest erosion risk.

Stream Temperature

Issues and Risks

Forest management activities can increase, maintain, or

decrease water temperature. Such changes can affect

drinking water quality (chapter 2) by altering dissolved

oxygen and survival rates of pathogens.

Findings from Studies

Surprisingly few studies have been published on the effects

of silvicultural practices on water temperature, and most of

these were conducted in the 1970’s (table 10.2). These

studies include harvesting with and without streamside

vegetation buffers. Several synthesis papers indicate that

few additional temperature studies have been conducted

(Beschta and others 1987, Binkley and Brown 1993a,

Swank and Johnson 1994).

Exposure of small streams to direct solar radiation is the

dominant process for stream temperature increases

(Tiedemann and others 1988). Other mechanisms including

increased air temperature, channel widening, soil water

temperature increases, and streamflow modification have

been proposed [Ice, in press (a)]. Small streams with smaller

surface areas may be more susceptible to heating, but

usually return to expected temperature within 500 feet [150

meters (m)] downstream [Andrus and Froehlich 1991; Ice,

in press (b)]. Maintaining shade in riparian zones can be

used to avoid most temperature increases in small streams.

As stream width increases, more of the water surface is

exposed to sunlight and the influence of riparian canopy on

stream temperature decreases.

Literature on the effects of timber harvesting on stream

temperatures (table 10.2) shows daily maximum stream

temperature increases from 1.2 to 7.2 oC in eastern forests

and 0.6 to 8 oC in western forests. The range in temperature

increases reflects a range in streamside vegetation buffers

from no buffer to a 100-m buffer. Changes in minimum

nighttime stream temperatures (during the winter or dormant

season) range from no change to <1 oC in the East and from

zero to <2 oC in the West.

Reliability and Limitations of Findings

Stream temperatures in small streams may increase after

timber harvesting when the streamside vegetation canopy is

removed. This effect can be mitigated by maintaining

streamside buffers. Several studies have reported temperature increases with streamside buffers, but increases are

much smaller than for fully exposed streams. The lack of
109

Chapter 10

Table 10.2—Effects of timber harvesting with and without streamside buffers on stream temperature

Maximum temperature

Location Treatment Temperature Change Measure Reference

- - - - Degree Celsius - - - -

East

Georgia Clearcut with buffer 25.0 3.9 Average daily Hewlett and Fortson 1982

Control 21.1

Maryland Riparian harvest 4.4 –7.6 Summer max. Corbett and Spencer 1975

Coweeta Hydrologic 100% clearcut with no buffer 21.7 3.4 Average daily Swift and Messer 1971

Laboratory, NC Control 18.3

Newark, NJ Riparian herbicide 3.3 Avg. summer max. Corbett and Heilman 1975

Fernow Experimental 95% clearcut with buffer removed 16.1 1.7 Average weekly Aubertin and Patric 1974

Forest, WV Control 14.4

Plot harvest 4.0 Summer max. Kochendorfer and Aubertin 1975

Hubbard Brook 100% clearcut with no buffer 20.0 4.0 Average daily Likens and others 1970

Experimental Control 16.0

Forest, NH

Pennsylvania State

Forest, PA Riparian harvest 3.9 Summer max. Lynch and others 1975

Leading Ridge, PA Control 19.4 1.2 Average daily Rishel and others 1982

44% clearcut with buffer 20.6

Control 17.8 7.2 Average daily Rishel and others 1982

85% clearcut with no buffer 25.0

West

Alsea, OR Control 12.2 10.0 Average daily Brown and Krygier 1970

85% clearcut with no buffer 22.2 16.0 Summer max.

Steamboat, OR Control 14.4 .6 Daily max. Brown and others 1971

Clearcut with buffer 15.0

Control 13.3 2.3 Daily max. Brown and others 1971

Clearcut with no buffer 15.6

British Columbia Clearcut .5–1.8 Average daily Holtby and Newcombe 1982

H.J. Andrews

Experimental

Forest, OR Clearcut 4.4–6.7 Daily max. Levno and Rothacher 1969

Coyote Creek, OR Clearcut 8.0 Daily max. Harr and others 1979
110

documentation on buffer characteristics makes extrapolation

difficult. Different measurements of stream temperature also

make direct comparisons difficult. Studies have reported

daily, monthly, or seasonal maxima or mean temperatures.

Within-stream temperature variability often is not considered in monitoring programs.

Attributes needed to estimate the contribution of forest

overstory to stream surface shade include stream width,

distance from vegetation to stream, stream orientation,

height and density of vegetation, crown or canopy measurement, latitude, date, and time (Quigley 1981).

A simpler model developed to predict the effect of

clearcutting on temperatures of small streams uses the

calculated heat load to the stream surface area (Brown

1970). This or similar models should be validated before

use. It would be difficult to suggest one streamwater buffer

model as suitable for all forest watersheds, but measurement

of the angular canopy density can determine the importance

of a buffer strip to prevent stream temperature increases

after timber harvesting. Angular canopy density is the

projection of the streamside vegetation canopy measured at

the angle above the horizon at which direct-beam solar

radiation passes through the canopy (Beschta and others

1987).

Generally, forest practices that open small stream channels

to direct solar radiation are the practices that increase stream

temperatures. Retention of streamside vegetation appears to

mitigate potential temperature changes, especially the

greater temperature changes. These principles are well

documented by research throughout the country. Streamside

canopy removal may also decrease winter streamwater

temperatures, since radiation losses may be increased. For

small streams, temperature returns to undisturbed levels

within a short distance downstream of where canopy shade

is reestablished.

Accurate stream temperature assessments vary from a single

instantaneous measurement to continuous measurement,

depending on the stream diel and seasonal variations.

Stream temperature data need to be evaluated over the long

term. Statistical methods include harmonic analysis, time

series, and trend analysis2 (Hostetler 1991, Limerinos

1978).

Research Needs

1. Stream temperature monitoring and reporting protocols

need to be developed.

2. The range or daily variation in temperature may increase

after removal of streamside vegetation. Research is

needed on these variations because they might affect

drinking water quality.

3. Timber harvesting with proper streamside vegetation

buffers should cause minimal stream temperature

changes. Stream buffers are defined by width only. More

studies need to be conducted investigating the efficiency

of different components of streamside canopy cover on

stream temperatures.

4. Stream temperature monitoring has tended to emphasize

physical measurements of temperature. Remote sensing of

stream temperature may provide more data on temperature changes over time and space.

5. Few water-quality related studies have assessed cumulative watershed effects. Temperature measurement studies

at different spatial scales need to be conducted. Long-term

temperature data are needed to place the potential effects

of changes in stream temperature in the context of global

or regional cycles of climate change or variability. Longterm records of stream temperature in undisturbed,

forested watersheds need to be collected.

Key Points

In general, removal of streamside vegetation cover has the

potential to increase streamwater temperatures during the

day in the summer. In certain settings, the vegetation

removal may allow for decreased nighttime temperatures,

especially in the winter. Temperature changes return to

pretreatment levels as the streamside vegetation reestablishes. Streamside vegetation to maintain a thermal cover

over the stream is key to maintaining stream temperatures at

existing levels.

Nutrients

Water from forested watersheds is typically lower in

nutrients than water that drains from other lands. Forest

management activities such as forest cutting and harvesting

may increase annual water yields (Bosch and Hewlett 1982,

Stednick 1996), interrupt the natural cycling of nutrients,

and increase nutrient concentrations in streamwaters.

Nitrogen and phosphorus cycles and their impacts on

drinking water quality are discussed in chapter 2. 2

 Stednick, J.D. 1999. Stream temperature trends in the New Alsea

watershed study. [55 p.]. Unpublished report. On file with: Department of

Earth Resources, Colorado State University, Fort Collins, CO 80523–1482.

Timber Management
111

Issues and Risks

Forest management activities, such as timber harvest and

fertilization, can increase nutrient concentrations in streams.

Findings from Studies

Nitrate nitrogen (NO3-N) concentrations are usually quite

low (0.002 to 1.0 milligrams per liter) in streams draining

undisturbed forest watersheds (Binkley and Brown 1993b).

Concentrations are low because nitrogen is used rapidly by

ecosystem biota and because nitrate formation (nitrification)

is relatively slow in forest soils. Slow rates of organic matter

decomposition, acid soil conditions common in forest

environments, and bacterial allelopathy all decrease rates of

nitrification. Organic matter and anaerobic conditions in

saturated riparian soils allow for denitrification, which is the

reduction of nitrate to nitrogen gas, which may be lost to the

atmosphere.

Throughout the United States, studies in many areas have

found that nutrient losses from silvicultural activities to be

minimal and water quality not degraded (Aubertin and

Patric 1974, Chamberlain and others 1991, Harr and

Fredriksen 1988, Hornbeck and Federer 1975, Martin and

others 1984, McClurkin and others 1987, Pierce and others

1972, Rense and others 1997, Sopper 1975, Swank 1988).

Nutrients contained in the organic matter in trees, litter, and

soils can be affected by various forest management practices. Cutting vegetation disrupts the processes that regulate

the nutrient cycle and may accelerate dissolved nutrient

leaching and loss via streamflow. Exposing sites to direct

sunlight may increase the rate of nitrogen mineralization.

Nutrients associated with eroded soil particles and sediment

may be lost from the site (Swank 1988). There is usually

minimal opportunity for a buildup of these nutrients in the

stream system after a timber harvest because of the normally

brief period of increased nutrient flux to the stream (Currier

1980). Other nutrients rarely cause water-quality problems,

and this discussion is limited to nitrogen and phosphorus.

Forest management activities such as harvesting or thinning

may interrupt nutrient cycles, and nutrients may be released

(Swank and Johnson 1994). Catchment studies have

produced a large body of information on streamwater

nutrient responses, particularly from clearcutting (table

10.3). Changes in streamwater nutrient concentrations vary

substantially among localities, even within a physiographic

region. In central and Southern Appalachian forests, nitratenitrogen, potassium (K+), and other constituents increased

after harvesting, but the changes were small and did not

affect downstream uses (Swank and others 1989).

Clear-cutting in northern hardwood forests may result in

large increases in concentrations of some nutrients

(Hornbeck and others 1987). Research on catchments has

identified some of the reasons for varied ecosystem response

to disturbance (Swank and Johnson 1994). Swank discusses

the long-term nitrate-nitrogen trends after harvest in chapter

11. In areas that are experiencing nitrogen saturation from

deposition of nitrogen compounds in air pollution, disturbances such as forest harvesting can produce increased

nitrate levels in streams and ground water (Fenn and others

1998). See chapter 3 for discussion of nitrogen-saturation

effects.

Soil development factors and forest management strategies

influence the rate of nutrient exports after timber harvesting

(Swank and Johnson 1994). The rotation length, the time

interval between timber harvests, is critical in determining

the sustainability of harvest. Nutrient loss by leaching to

streams is usually minor compared to the nutrient loss by

biomass removal (Clayton and Kennedy 1985, Federer and

others 1989, Johnson and others 1988, Mann and others

1988, Martin and Harr 1989). Nutrient loss differences are

also observed between whole tree, saw log, or bole-only

harvesting.

Phosphorus (P) occurs in several forms in surface water

including the dissolved forms of orthophosphates and

dissolved complex organics and in particulate forms

(organic and inorganic) [Ice, in press (b)]. Phosphorus

sources come from dry deposition (dust), wet deposition,

and geologic weathering. Geology is a key factor in phosphorus concentrations from forests. Forest watersheds with

more easily weathered rock, such as sedimentary or volcanic

tuff and breccia, have higher instream concentrations than

watershed with resistant rock, such as intrusive igneous.

Dissolved phosphorus is probably one of the least responsive water-quality constituents to forest management.

Total phosphorus is strongly associated with soil particles or

suspended sediment. Practices that increase or reduce

sediment have similar effects on total phosphorus [Ice, in

press (b)].

In general, nutrient mobility from disturbed forests follows

the order: nitrogen > potassium > calcium and magnesium

> phosphorus. Thus, forest harvesting or other disturbances,

such as fire, will generally produce larger differences in

nitrogen concentrations in streamwater than other constituents. Possible exceptions are the loss of calcium and

potassium documented in the Northeastern United States

where precipitation inputs had greater acidity from fossil

fuel combustion (Federer and others 1989).

Chapter 10
112

Table 10.3—Effects of clearcutting with and without buffers on mean annual nitrate-nitrogen, ammonium-nitrogen, and

total-phosphorus concentrations

Mean concentration

Location Treatment NO3-N NH4-N Total Pa Reference

- - - - - - - - - - - - - Milligrams per liter - - - - - - - - - - - - -

East

Marcell Experimental 74% clearcut 0.16 0.55 Verry 1972

Forest, MN Control .12 .41

Hubbard Brook

Experimental

Forest, NH

WS2 100% cut and

herbicide 8.67 – 11.94 .04 – .05 0.002 Likens and others 1970

WS4 33% strip cut .19 – .20

WS6 Control .16 – .29 .05 – .09 .001

White Mountain, NH

Seven catchments Control .02 – .81 Pierce and others 1972

Clearcut 1.31 – 3.84 .01 – .02

Upper Mill Brook Control .23 – .27 .02 – .03 Stuart and Dunshie 1976

Clearcut .23 – .96

Leading Ridge, PA

LR2 100% clearcut

and herbicide .10 – 8.4 Corbett and others 1975

LR1 Control .02 – .04

Fernow Experimental

Forest, WV

WS3 100% clearcut .18 – .49 .14 – .35 .04 – .07 Aubertin and Patric 1972, 1974

WS4 Control .10 – .32 .13 – .48 .02 – .04

Coweeta Hydrologic

Laboratory, NC

WS2 Control .004 .002 .006 Douglass and Swank 1975

WS28 100% clearcut .094 .003 .004

West

H.J. Andrews Control .020 – .200 .016 – .032 Fredriksen and others 1975

Experimental 100% clearcut .001 – .010 .024 – .039

Forest, OR

Bull Run, OR 25% clearcut .002 – .093 .001 – .005 .011 – .032 Fredriksen 1971

Control .002 – .013 .002 – .005 .014 – .040

Coyote Creek, OR 100% clearcut .001 – .275 .001 – .018 .062 – .100 Harr and others 1979

Control .001 – .005 .001 – .014 .036 – .060 Adams and Stack 1989

Chicken Creek, UT 13% clearcut .025 Johnston 1984

Control .008

Alsea, OR 85% clearcut .19 – .44 Brown and others 1973

Control 1.18 – 1.21

Priest River, ID Control .20 Snyder and others 1975

100% clearcut .18

Fraser Experimental 33% clearcut .06 Stottlemeyer 1987

Forest, CO Control .006

Beaver Creek, AZ Control .010 Ryan as cited by

Clearcut .220 Binkley and Brown 1993b

LR = Leading Ridge; NO3

-N = nitrate-nitrogen; NH4-N = ammonium-nitrogen; total P = total phosphorus; WS = watershed. a Blank columns represent no data collected.

Timber Management
113

Reliability and Limitations of Findings

Research has documented that timber harvesting may

increase nitrate concentrations in soil water and streams.

This finding is generally accepted without controversy.

Soluble phosphorus concentrations are essentially unaffected by timber harvesting activities. Total phosphorus

concentrations are closely linked to sediment concentrations. Some forest types in the United States have few

studies investigating the influence of forest practices on

water quality. The rather consistent streamwater chemistry

response to timber harvesting allows response extrapolation.

However, an often erroneously cited study as an example of

timber harvesting effects on water quality is an early

Hubbard Brook study (Likens and others 1970). In this

study, vegetation was cut, left onsite, and sprayed with a

general herbicide for 3 years to kill any plant regeneration to

research nutrient cycling processes. Nutrient concentrations,

particularly nitrate, increased significantly. This watershed

treatment was not representative of timber harvest and does

not represent the effects of a typical timber harvest on water

quality.

If vegetation is quickly reestablished, nutrient exports are

short-lived and usually do not represent a threat to water

quality or site productivity. There are a couple of possible

exceptions. Nitrogen deposition can accumulate in forest

soils over time, especially in areas with air-quality concerns

(Riggan and others 1985, Silsbee and Larson 1982). If

timber harvesting occurs in these areas, mobilization of

accumulated soil nitrogen may result in higher nitrate

concentrations and outputs in the streamwater (see

chapter 3).

In the Pacific Northwest, water-quality samples from

streams in forests with nitrogen fixing alder (Alnus spp.)

may have higher nitrate concentrations than streams without

alder (Binkley and Brown 1993b, Miller and Newton 1983).

Since nitrogen is being added to the site by fixation, losses

in site productivity are not a concern, but nitrate concentrations may be high enough to affect downstream uses.

Forest harvesting practices that minimize site disturbance

and quickly establish new stands seem to minimize any

potential water-quality effects. Streamside vegetation

buffers are effective for sediment removal and nutrient

removal.

Research Need

Soil water usually has higher nutrient concentrations than

surface or streamwater. Changes in water chemistry at large

scales (watershed to landscape) need to be evaluated,

especially in the context of multiple land-use activities in

time and space for cumulative watershed effects.

Key Point

Timber harvesting may increase nutrient concentrations

in streams, especially nitrate, but any increase is usually

short-lived. Watershed studies show that nutrient concentrations in soil water may be higher than concentrations in

surface water suggesting that other water dilutes off-site

concentrations.

Fertilizer

Urea fertilizer is highly soluble in water and readily moves

into the forest floor and soil with any appreciable amount of

precipitation. Under normal conditions, urea is rapidly

hydrolyzed (4 to 7 days) to the ammonium ion (NH4-N).

When moisture is limited, urea may be slowly hydrolyzed

on the forest floor. Rather than moving into the soil as

ammonium, the increased soil surface pH favors formation

of ammonia (NH3-N), which is lost by volatilization.

Volatilization losses may be significant. Fertilizer usually is

applied in the spring or fall to take advantage of seasonal

precipitation.

Fertilizers may enter surface water by several routes. Direct

application of chemicals to exposed surface water is the

most significant. Identification of surface water bodies prior

to the application essentially eliminates this entry mode.

When fertilizers are volatilized, ammonia absorption by

surface water is minimal (U.S. Department of Agriculture,

Forest Service 1980).

Issues and Risks

The issues and risks associated with fertilizer application are

essentially the same as described in the Nutrients section,

except if inadvertently applied to streams.

Findings from Studies

The reported effects of forest fertilization on water quality,

particularly nutrient concentrations in streams, are variable3

(reviews by Binkley and Brown 1993b, Binkley and others

1999, Bisson and others 1992, Fredriksen and others 1975).

Nutrient retention by forest soils is excellent. Nutrient

3

 Stephens, R. 1975. Effects of forest fertilization in small streams on the

Olympic National Forest, fall 1975. Unpublished report. 40 p. On file with:

USDA Forest Service, Olympia Forestry Sciences Laboratory, 3625 93rd

Avenue, South, Olympia, WA 98512 .

Chapter 10
114

concentrations in surface waters after forest fertilization are

usually low (table 10.4). Exceptions may occur in areas

experiencing nitrogen saturation from air pollution inputs.

For example, Fernow Experimental Forest, WV, a site that

shows signs of nitrogen saturation (Fenn and others 1998),

experienced high streamwater nitrate response to nitrogen

fertilization (table 10.4). Ammonium-nitrogen and phosphorus are very reactive with forest soils and are retained on

site. Ammonium-nitrogen concentration may increase in

surface water as a result of direct fertilizer application to

open water. Ammonium-nitrogen concentrations, however,

are rapidly reduced through aquatic organism uptake and

stream sediment sorption. See chapter 3 for discussion of

surface and ground water responses to nitrogen additions in

nitrogen-saturated watersheds.

Table 10.4—Effects of forest fertilization on maximum streamwater ammonium-nitrogen and nitrate-nitrogen

concentrations

Location Treatment NH4-N NO3-N Reference

Lbs/ac - - - Milligrams per liter - - - -

East

Fernow Experimental

Forest, WV 230 0.8 19.8 Aubertin and others 1973

West

Coyote Creek, OR 200 .04 .17 Fredriksen and others 1975

Olympic National

Forest, WA 200 .02– .55 .07–3.85 Stephens 1975a

200 .04 .121 Moore 1975

Entiat Experimental

Forest, WA 48 <.02 .210 Klock 1971

50 .068 Tiedemann and Klock 1973

Mitkof Island, AK 187 .003 2.36 Meehan and others 1975

Siuslaw River, OR 200 .49 7.6 Burrough and Froehlich 1972

Cascade Mountains, OR 200 <.01 <.25 Malueg and others 1972

Lake Chelan, WA 70 .011 .510 Tiedemann 1973

South Umpqua River, OR 200 .048 .177 Moore 1971

Ludwig Creek, WA 178 .004 2.7 Bisson and others 1992

NH4-N = ammonium nitrogen; NO3-N = nitrate nitrogen. a Stephens, R. 1975. Effects of forest fertilization in small streams on the Olympic National Forest, fall 1975. 40 p. Unpublished report. On file

with: Olympic National Forest, 1835 Black Lake Boulevard, SW, Olympia, WA 98512.

Nitrate-nitrogen concentrations measured in surface water

usually peak 2 to 4 days after fertilizer application (U.S

Department of Agriculture, Forest Service 1980). The

magnitude of the peak concentration may depend on the

presence and width of streamside buffers and the density of

small feeders and tributaries to the streams. Peak nitratenitrogen concentrations usually decrease rapidly but may

remain above pretreatment levels for 6 to 8 weeks. Winter

storms may also result in peak nitrate-nitrogen concentrations, but these peaks usually decrease over successive

storms, and concentrations decrease quickly between

storms.

Timber Management
115

Reliability and Limitations of Findings

Relatively few studies have been published on the effect of

forest fertilization on water quality, but results generally are

consistent and suggest that concentrations of ammoniumnitrogen and phosphorus do not increase after fertilization

(NCASI 1999b). Nitrate-nitrogen concentrations may

increase, but increases are short-lived. Publications reviewed here suggest minimal water-quality changes under

most conditions and appear universally applicable.

Streamwater responses to fertilizer application are well

understood and may be extrapolated. An exception to this

generalization may be areas showing signs of nitrogen

saturation. Nitrogen fertilization in these areas may increase

stream nitrate.

Forest fertilization may increase nitrate-nitrogen concentrations by direct application of fertilizer to the stream or by a

runoff-generating precipitation event after application.

Careful delineation of application areas will avoid direct

stream inputs. Fertilizer application timing with respect to

seasonal precipitation or storm events minimizes fertilizer

affects on water quality.

Research Needs

1. Streamside vegetation buffers or management zones are

usually prescribed as a width. We need to know what

specific components or processes in these streamside

areas would minimize the movement of fertilizers into

surface water.

2. Recent research identified certain bedrock materials as

significant sources of nitrogen. Heretofore, geologic

materials were not considered significant sources of

nitrogen. How common are these materials?

3. What are the effects of repeated fertilizer applications in

short-rotation forest plantations on water quality?

4. Response of stream nitrate to fertilization in areas

experiencing nitrogen saturation is poorly understood

and needs more study.

Key Points

Application of nitrogen or phosphorus fertilizers will not

adversely affect surface waters including drinking waters,

when the fertilizer is applied at a rate and time when the

vegetation can use it. Fertilizer application should be timed

to avoid rainy periods if fertilizer might be moved directly

to surface waters. Streamside vegetation is effective in

nutrient removal. Any increase in nutrient concentrations

from fertilizer applications is usually short-lived and should

not affect downstream uses.

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Malueg, K.W.; Powers, C.F.; Krawczyk, D.F. 1972. Effects of aerial

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Mann, L.K.; Johnson, D.W.; West, D.C. [and others]. 1988. Effects of

whole -tree and stem only clearcutting on post harvest hydrologic losses,

nutrient capital and regrowth. Forest Science. 34: 412–428.

Martin, C.W.; Harr, R.D. 1989. Logging of mature Douglas-fir in western

Oregon has little effect on nutrient output budgets. Canadian Journal of

Forest Research. 19: 35–43.

Martin, C.W.; Noel, D.S.; Federer, C.A. 1984. Effects of forest

clearcutting in New England on stream chemistry. Journal of Environmental Quality. 13: 204–210.

McClurkin, D.C.; Duffy, P.D.; Ursic, S.J.; Nelson, N.S. 1985. Water

quality effects of clearcutting Upper Coastal Plain loblolly pine

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McClurkin, D.C.; Duffy, P.D.; Ursic, S.J.; Nelson, N.S. 1987. Changes

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20-year-old pine. Journal of Environmental Quality. 16: 237–241.

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ARS–40. Beltsville, MD: U.S. Department of Agriculture, Agricultural

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Miller, E.L.; Beasley, R.S.; Lawson, E.R. 1988. Forest harvest and site

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annual meeting of the western reforestation coordination committee;

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OR: Western Forestry and Conservation Associates: 28–35.

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OR: U.S. Department of Agriculture, Forest Service, Pacific Northwest

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53 p.

Patric, J.H.; Evans, J.O.; Helvey, J.D. 1984. Summary of sediment yield

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Chapter 10
120

Chapter 11

Forest Succession

Wayne Swank1

Introduction

The effects of forest management activities on water quality

are generally of the greatest magnitude in the first several

years after disturbance. However, during long-term succession and regrowth of forest ecosystems, changes in physical,

chemical, and biological parameters of streams may occur.

Nutrients

Issues and Risks

After a forest disturbance such as harvesting or fire, nutrient

levels in streams may be elevated during early successional

stages until the forest matures (see chapter 10). Nitrate

concentrations can be elevated for a few to many years

depending upon whether the watershed is nitrogen limited

or saturated (see chapter 3 for discussion of nitrogen

saturation).

Findings from Studies

Changes in stream inorganic chemistry and sediment yield

were observed over a 20-year period after clearcutting by

cable logging of a 146-acre [58-hectare (ha)] Southern

Appalachian watershed (Swank and others, in press). Stream

nutrient concentrations and fluxes showed small increases

after harvest, and responses were largest the third year after

treatment. Nitrate-nitrogen (NO3

-1) was an exception. The

initial increase in nitrate was from <0.1 milligrams (mg) per

liter to 0.8 mg per liter (fig. 11.1) and increased net nitrogen

export of 1.16 pounds per acre [1.3 kilograms (kg) per

hectare] the third year after harvest. However, later in

succession (15 to 20 years), nitrate concentrations exceeded

values observed the first several years after clearcutting.

This response is partially attributed to reduction in nitrate

uptake due to vegetation mortality, changes in species

composition, and nitrogen release from decomposition of

woody plants.

Other long-term research in eastern forested watersheds

(Edwards and Helvey 1991, Swank and Vose 1997) shows

that as forests mature, less nitrogen is retained in the

watershed and stream nitrate concentrations increase. These

long-term studies support findings of shorter term stream

chemistry surveys. A survey of streamwater chemistry in 57

watersheds along successional and elevational gradients was

conducted in the White Mountains of New Hampshire

(Vitousek 1977). Differences in successional status among

watersheds were found to be important in controlling nitrate

and potassium concentration. Streams draining old-aged

forests had higher concentrations of nitrate, potassium, and

other solutes than did streams draining intermediate-aged

forests at the same elevation. Spruce-fir (Picea spp.-Abies

spp.) watersheds with no record of logging had streamwater

nitrate concentrations of about 3 mg per liter, while sprucefir watersheds logged 30 years previously had nitrate

concentrations <0.5 mg per liter.

Another survey of 38 streams draining partially or entirely

clearcut watersheds was conducted in New England—

(Martin and others 1985) on northern hardwood sites in

New Hampshire, Maine, and Vermont; in central hardwood

forests in Connecticut; and in coniferous forests in Maine

and Vermont. Streams draining watersheds that had been

partially or entirely clearcut in the previous 2 years were

selected. There were no apparent changes in stream nutrient

concentrations from many of the ecosystems, and the largest

concentration increases were for nitrate, calcium, and

potassium in northern hardwoods of New Hampshire.

Inorganic nitrogen (nitrate plus ammonium) increased to an

average of 2 mg per liter (Martin and others 1985). However, elevated solute concentrations appear to be short-lived,

even in streams draining successional northern hardwood

forests in New Hampshire (Hornbeck and others 1987).

Moreover, early stream chemistry changes after clearcutting

were considered insufficient to cause concern for public

water supplies or for downstream nutrient loading (Martin

and others 1985).

In the Pacific Northwest, forest-successional stage is not

always a good predictor of nitrate concentration in

streamwater. For example, at the H.J. Andrews Experimental Forest in Oregon, forest harvest increased annual nitrate 1

 Retired Project Leader, USDA Forest Service, Southern Research Station,

Otto, NC.
121

concentration from predisturbance levels of 0.001 mg per

liter to 0.036 mg per liter (Martin and Harr 1989), but nitrate

concentration returned to predisturbance levels within 6

years. Further, a 20-year postdisturbance record from a pair

of treated and untreated watersheds at the experimental

forest suggests that nitrate concentrations in streamwater

remain very low in both watersheds once the clearcut

watershed recovers from the immediate effects of disturbance.2 At the H.J. Andrews Experimental Forest, the

ecosystem is highly nitrogen-limited, and vegetation imprint

on nitrogen fluxes may be overridden by rapid immobilization of any available nitrogen by soil microbiota.

An extensive synoptic water-quality assessment was

conducted on numerous streams in the Great Smoky

Mountain National Park in the Southern Appalachian

Mountains (Flum and Nodvin 1995, Silsbee and Larson

1982). Concentrations of nitrate in streams draining watersheds that had been logged prior to park establishment were

significantly lower (one-half) than the nitrate concentrations

in unlogged watersheds at similar elevations.

The magnitude of stream nitrate concentrations associated

with long-term forest succession depends on a number of

factors, such as levels of atmospheric nitrogen deposition,

the type and rapidity of forest regrowth, soil microbial

activity, and soil physiochemical reactions. Stream nitrate

levels rarely exceed 5 mg per liter and are below current

drinking water standards. The nitrate, however, may

contribute to stream acidification, particularly during spring

snowmelt when nitrate concentrations peak in the Northeastern United States (Murdock and Stoddard 1992).

Reliability and Limitations of Findings

Existing evidence for changes in stream chemistry with

forest succession is based upon well-established programs

of long-term research and is quite reliable. However,

findings are limited in scope to select forest ecosystems in

the United States.

Limited evidence indicates that stream nitrate concentrations

for older hardwood forests of the southern and central

Appalachian regions are higher than for younger successional forests. However, site-specific research shows that

nitrate levels can vary substantially even during early

succession (first 20 years), although the general applicability

of findings is unknown. Assessments of nitrate levels in

streams draining successional forests in New England show

mixed responses and appear to be ecosystem specific. Very

limited information on stream nitrate is available for successional forests in the Pacific Northwest. Current findings

2

 Personal communication. 1999. Kristin Vanderbilt, Graduate Student.

Oregon State University, Corvallis, OR.

WS 2

WS 7

1.00

.90

.75

.60

.45

.30

.15

0

1975 1980 1990 1985 1995

Year

Road construction

Road seeding and fertilization

Cutting and logging

Site preparation

Stream NO3-1 concentration (mg/L)

Figure 11.1—Mean monthly concentrations (flow weighted) of nitrate (NO3

-1) in streamwater of a clearcut, cable-logged,

hardwood-covered watershed (WS7) and an adjacent watershed (WS2) during calibration, treatment activities, and

postharvest period, Coweeta Hydrologic Laboratory, North Carolina.

Chapter 11
122

indicate that elevated nitrate concentrations following clearcutting are short-lived and return to predisturbance levels

early in succession.

Research Needs

1. Long-term assessments of stream chemistry changes

associated with forest succession are lacking for most

major forest ecosystems in the United States. From a

public drinking water perspective, synoptic stream

nutrient surveys across a range of forest types and stand

ages with known disturbance histories would greatly

enhance planning information for managers.

2. There is a large knowledge gap in nutrient concentration

changes associated with storm runoff events. Such

information is most important where water supplies are

derived from forested headwaters with rapid streamflow

responses to precipitation, e.g., watersheds with shallow

soils, steep slopes, intense rainfall, and rapid snowmelt.

Sediment

Issues and Risks

Stream sediment may also exhibit long-term dynamics after

forest disturbance. Logging roads associated with harvesting

activities are frequently the major source of sediment to

streams and are a potential legacy to consider when evaluating sources of sediment in drinking water (see chapters 3, 9).

Findings from Studies

A synthesis of long-term sediment yield responses following

forest watershed disturbances is provided by Bunte and

MacDonald (1999). Based on studies in Oregon and New

Hampshire, they identify three kinds of potential responses

in postdisturbance sediment yields:

1. Sediment yields remain high for a number of years after

disturbance due to a large sediment pulse to the stream from

a storm or other disturbance. That is, sediment from upstream storage areas or destabilized hillslopes and channels continues to be released;

2. Sediment yields decline below average annual yields

after disturbance when sediment storage is depleted by a

major sediment transport event; and

3. Sediment yields rapidly return to predisturbance conditions because excess material has moved through the

system.

Recent findings in the Southern Appalachians provide an

example of the first type of response where sediment yield

remains high for a number of years during forest succession

(Swank and others, in press). A cable-logged, clearcut

watershed required only three contour access roads because

logs could be yarded 1,000 feet (305 meters) with the cable

system. Record storms (15 inches or 38 centimeters) in the

last 2 weeks of May 1976, prior to grass establishment,

eroded both unstable soil and hydroseeded materials from

the roads. Roads were the source of elevated sediment yield

as illustrated by soil loss measured at a gaging station in the

stream immediately below a road crossing in the middle of

the catchment (fig.11.2A). In those 2 weeks of May,

sediment yield was nearly 55 tons [50 metric tonnes (Mg)]

from 0.21 acre (0.085 ha) of road contributing area (roadbed, cut, and fill). In the ensuing period of road stabilization

and minimum use (June to December 1976), soil loss was

low but accelerated again briefly during the peak of logging

activities (fig. 11.2A). In the next year, soil loss below the

road declined to baseline levels.

The pattern of sediment yield at the base of the second-order

stream (fig. 11.2B, gaging site) draining the watershed was

different from the pattern of sediment loss from the roads.

Following an initial pulse of sediment export from the

watershed, sediment yield remained substantially elevated

during and after logging. In the 3-year period between

1977–80, the cumulative increase in sediment yield was 240

tons (218 Mg) (fig. 11.2B). During the next 10 years,

sediment yield declined with a cumulative increase in export

of 240 tons (218 Mg). The rate of sediment yield over the 5-

to 15-year period after disturbance was about 300 lb per acre

per year (336 kg per hectare per year), or 50 percent above

pretreatment levels. The long-term sediment yield data

illustrate a lag or delay between pulsed sediment inputs to a

stream and the routing of sediments through the stream

channels. In the absence of significant additional sources of

sediment to streams on the watershed, annual sediment yield

at the base of the watershed was still substantially above

predisturbance levels at least 15 years later. Thus, there

appears to be a continual release of sediment from upstream

storage that was primarily deposited from road crossings of

streams during exceptionally severe storms.

Reliability and Limitations of Findings

Few studies have documented the long-term effects of

management practices on sediment yield. As pointed out in

chapter 10, increases in sediment yields from timber

management activities are typically considered to be shortlived. However, unique conditions during management can

lead to elevated stream sediment later in forest succession.

The importance of this process is site-specific and requires

Forest Succession
123

Cumulative sediment yield

(tons)

1975

0

1977 1979 1981 1983 1985 1987 1989

100

200

300

400

500

600

700

Date

(B)

Measured

Predicted

that each stream be evaluated to assess the legacy of past

management practices on current levels of stream

sedimentation.

Research Need

Recommendations for future research related to this topic

are given in chapter 10.

Key Points

In the long term, forest harvesting practices alone may have

little deleterious impact on stream sediment and chemistry,

which are of primary concern in drinking water. However,

other past and present land uses affect present sediment and

nitrate concentrations in streams. Sediment and nitrate

yields associated with early successional development of

forest may be in addition to yields from other past and

present land uses. It is important to consider successional

impacts along with the cumulative impacts of other past and

present land uses across the landscape when assessing

impacts of land management on drinking water sources.

Literature Cited

Bunte, K.; MacDonald, L.H. 1999. Scale considerations and the

detectability of sedimentary cumulative watershed effects. Tech. Bull.

776. Research Triangle Park, NC: National Council for Air and Stream

Improvement. 327 p.

Edwards, P.J.; Helvey, J.D. 1991. Long-term ionic increases from a

control Appalachian forested watershed. Journal of Environmental

Quality. 20: 250–255.

Flum, T.; Nodvin, S.C. 1995. Factors affecting streamwater chemistry in

the Great Smoky Mountains, U.S.A. Water, Air and Soil Pollution. 85:

1707–1712.

Hornbeck, J.W.; Martin, C.W.; Pierce, R.S. [and others]. 1987. The

northern hardwood forest ecosystem: ten years of recovery from

clearcutting. Res. Pap. NE–596. Broomall, PA: U.S. Department of

Agriculture, Forest Service, Northeastern Forest Experiment Station.

30 p.

Martin, C.W.; Harr, R.D. 1989. Logging of mature Douglas-fir in western

Oregon has little effect on nutrient output budgets. Canadian Journal of

Forest Research. 19: 35–43.

Martin, C.W.; Noel, D.S.; Federer, C.A. 1985. Clearcutting and the

biogeochemistry of streamwater in New England. Journal of Forestry.

83(11): 686–689.

Murdock, P.S.; Stoddard, J.L. 1992. The role of nitrate in the acidification of streams in the Catskill Mountains of New York. Water Resources

Research. 28(10): 2707–2720.

Silsbee, D.G.; Larson, G.L. 1982. Water quality of streams in the Great

Smoky Mountains National Park. Hydrobiologia. 89: 97–115.

Swank, W.T.; Vose, J.M. 1997. Long-term nitrogen dynamics of Coweeta

forested watersheds in the Southeastern United States of America.

Global Biogeochemical Cycles. 11(4): 657–671.

Swank, W.T.; Vose, J.M.; Elliott, K.J. [In press]. Long-term hydrologic

and water quality responses following commercial clearcutting of mixed

hardwoods on a Southern Appalachian catchment. Forest Ecology and

Management.

Vitousek, P.M. 1977. The regulation of element concentrations in

mountain streams in the Northeastern United States. Ecological

Monographs. 47: 65–87.

Figure 11.2—Cumulative sediment yield measured on a clearcut, cablelogged, hardwood-covered watershed: (A) in one of the first-order streams

below a logging road during the first 32 months after treatment and (B) in

the ponding basin of the second-order stream at the gaging site during 15

years after treatment. Predicted values are based on pretreatment calibration

of sediment yield with an adjacent control watershed, Coweeta Hydrologic

Laboratory, North Carolina.

Apr Nov July Apr Nov July Apr Nov July

1976 1977 1978

Logging activities

Road construction

Cumulative sediment yield

(tons)

(A) 65

50

40

30

20

0

Two major storm events

Chapter 11
124

Chapter 12

Fire Management

Johanna D. Landsberg and Arthur R. Tiedemann1

Introduction

The effect of wildfire on drinking water was graphically

demonstrated when the Buffalo Creek fire in Colorado in

1996 was followed by heavy rains, forcing municipal water

supplies to shut off, one of Denver’s water treatment plants

to close, months to be spent cleaning a water-supply

reservoir, and the Coors Brewing Company to bring in water

by truck (Illg and Illg 1997).

Fire, both wild and prescribed, has the potential to alter

physical, chemical, and biological properties of surface

water that originates from burned wildland areas. Nonpointsource pollution from wildland after fire can impair the

suitability of water for drinking and other purposes. New

plans for widespread use of prescribed fire to solve forest

health problems create an urgent need to fully understand

the water-quality consequences of increasing the occurrence

of fire. Fire management activities (like retardant application, fireline construction, and postfire rehabilitation) also

have potential effects on water quality.

The most important effects of fire on drinking water source

quality include sediment and turbidity or both, water

temperature, and increased nutrients in streamflow. In this

chapter, we review results of research on the response of the

above water-quality variables to fire, fire management

activities, and fire rehabilitation measures. Much of the

information comes from reports on wildfires. We would

expect the magnitude of streamwater-quality changes after

prescribed fire to be less than those observed after wildfires

and some broadcast slash burns. It is unlikely that prescribed

fire would consume as much forest floor and understory, or

kill as much overstory, as would a wildfire because prescribed fires are usually conducted under conditions

deliberately chosen to produce burns of low severity.

Sediment and Turbidity

Issues and Risks

Suspended sediment is the major nonpoint-source pollution

problem in forests (Society of American Foresters 1995).

Beschta (1990) reported that sediment and turbidity are the

most significant water-quality responses associated with

fire. Turbidity has no direct health effects but can interfere

with disinfection and provides a medium for microbial

growth. Thus, it may indicate the presence of microbes (U.S.

EPA 1999). See chapter 2 for more discussion on the effects

of sediment on drinking water.

Findings from Studies

To understand research findings about sediment production

and its impacts, one must be familiar with the units of

measurement in which sediment is reported. Suspended

sediment is particles carried in suspension and is measured

by filtering and drying a known volume of water. Suspended

sediment is expressed in parts per million (ppm), or as

turbidity in nephelometric turbidity units (NTU’s), which is

a measure of the cloudiness of the water. These methods

measure different characteristics of water, and it is difficult

to correlate the results of one method with results of the

other. The standard turbidity method (U.S. EPA 1999) uses

NTU’s. We found only two studies of fire effects that

reported results in NTU’s [equivalent to Jackson turbidity

units (JTU’s)] from American Public Health Association

(1976) (table 12.1); all others reported sediment in parts per

million (table 12.2). Beschta (1980) found that a relationship

between suspended sediment and turbidity can be established but that the relationship differs significantly among

watersheds. He suggested that the relationship must be

established on a watershed-by-watershed basis. Recognizing

this difficulty, Helvey and others (1985) determined the

relationship between sediment in parts per million and

turbidity in NTU’s for three catchments in northcentral

Washington and found the relationship to be strong (Helvey

and others 1985). With this strong relationship and the

equations developed, sediment measurements, in parts per

million, can be converted to turbidity measurements, in 1

 Research Fire Ecologist/Soil Scientist and Scientist Emeritus, USDA

Forest Service, Pacific Northwest Research Station, Forestry Sciences

Laboratory, Wenatchee, WA, respectively.
125

NTU’s (fig. 12.1). This relationship has not been tested in

other geographic areas or plant community types, so caution

is advised when applying it beyond its original limits. It

does point out the need to use a standard method or to

establish relations between suspended sediment and

turbidity for each watershed or stream system in question.

Our interest here is on the effects of fire on sediment

measured in NTU’s. Wright and others (1976, 1982) found

that slope plays an important role in the amount of turbidity

in streamflow after broadcast burning oak-juniper (Quercus

spp.-Juniperus spp.) watersheds in central Texas. Turbidity

changes (table 12.1) after burning were most pronounced in

the steepest watersheds, with levels reaching 230 JTU’s.

Studies of suspended sediment (table 12.2) show that the

range of the prefire or control values is 1 to 26 ppm. Values

obtained after fires indicate that fire has a profound effect on

sediment movement.

Table 12.1—Water turbidity, in Jackson turbidity units (equivalent to nephelometric turbidity units), after fire

alone or in combination with other treatments

Pretreatment

Treatment Habitat Location or control Posttreatment Reference

- - Jackson turbidity units - -

Prescribed fire, Juniper Central Texas Wright and others 1976

pile, and burn 3 to 4% slope 12 12

8 to 20% slope 20 53

37 to 61% slope 12 132

Pile and burn Juniper Central Texas 12 162 Wright and others 1982

Pile, burn, and 12 72

seed

Table 12.2—Suspended sediment concentration in streamflow after fire alone or in combination with

other treatments

Pretreatment

Treatment Habitat Location or control Posttreatment Reference

- - - - - Parts per million - - - - -

Wildfire Taiga Interior Alaska 3.7 –10.6 2.6 – 6.0 Lotspeich and others 1970

Clearcut, slash

broadcast burned Douglas-fir Western Oregon 2 56 –150 Fredriksen 1971

Wildfire Ponderosa pine, Eastern

 Douglas-fir Washington Not known 1,200a Helvey 1980

Pile, burn Juniper Central Texas 1.1 3.7 Wright and others 1982

Pile, burn, and

seed 1.0 3.7

Prescribed fire Loblolly pine Upper Piedmont,

 plantation South Carolina 26 33 Douglass and Van Lear 1983

Wildfire Lodgepole pine, Glacier National < 3 15 – 32 Hauer and Spencer 1998

 Douglas-fir, Park, MT

 ponderosa pine,

 western larch

a

 Maximum value attained.

Chapter 12
126

Sediment yield has been measured in pounds per acre per

year in many studies because of the concern for soil loss

after fire. Sediment yield varies widely as a consequence of

fire or forest harvest and fire (table 12.3). This variability

reflects numerous interacting factors: geology, soil, slope,

vegetation, fire characteristics, treatment combinations,

weather patterns, and climate.

In the research we reviewed, sediment yield from pretreatment or control areas ranged from as low as 3 pounds per

acre per year [3.36 kilograms (kg) per hectare per year] to as

high as 12,500 pounds per acre per year [14 metric tonnes

(Mg) per hectare per year] (table 12.3). Postburn sediment

yield ranged from as low as 12 pounds per acre per year

(13.5 kg per hectare per year) to as high as 98,160 pounds

per acre per year (110 Mg per hectare per year). The lower

values generally were associated with flatter land and lower

severity fires. The higher values resulted from more severe

fires on steeper slopes and from fires on areas with soils

formed from decomposing granite, which erode readily.

When fire is used to convert brush to grass, it can have an

unintended side effect: mass wasting, which can affect water

quality. Work in California established the susceptibility of

steep slopes to mass soil movement following conversion of

brush to grass (Bailey and Rice 1969). These mass soil

movements produce long-lasting changes. In one study,

these same effects occurred on steep, forested slopes;

especially after severe fires (Robichaud and Waldrop 1994)

(table 12.3). These sediment yields are sufficient to generate

concern about water turbidity, which was not measured

directly.

Burned areas are sometimes seeded to rapidly establish

plants or are given other treatments to quickly stabilize the

soil. Following severe wildfire, the Forest Service and other

land managers sometimes implement Burn Area Emergency

Rehabilitation (BAER) treatments to reduce the risk of high

runoff and sediment flows to vulnerable installations

downstream such as drinking water intakes and reservoirs.

In a review of literature and monitoring reports, Robichaud

and others (in press) found that the effectiveness of the most

widely used BAER practice, contour-felled log barriers, had

not been systematically studied. The second most used

BAER practice, postfire broadcast seeding with grasses, has

been studied and the majority of studies found that this

treatment did not significantly reduce erosion during the

critical first 2 years after fire (Robichaud and others 2000).

Effectiveness of contour felling has not been tested, and

reseeding with grasses is not a reliable technique for erosion

control after severe wildfire. Additionally, when an area is

seeded with nonnative grass species, native plant species

may be effectively excluded leading to questions about

long-term stability (Tiedemann and Klock 1976).

Firelines, particularly those that are created by bulldozers,

are important potential sources of suspended sediment and

turbidity in streams for several reasons. First, some firelines

are constructed in urgent circumstances, without adequate

time to consider stream protection. Thus, they may provide

direct channels for sediment into streams. Second, firelines

may be difficult to stabilize with vegetation because much

of the nutrient-rich surface soil is cast aside. Hence, they are

likely to be slow to revegetate with perennial vegetation.

Information on revegetating and stabilizing firelines is very

limited. Two studies found application of seed and fertilizer

is an effective way to protect firelines (Klock and others

1975, Tiedemann and Driver 1983). Klock and others (1975)

demonstrated that seeding firelines with several species of

introduced and native grasses produced up to 85 percent

foliar cover within 2 years. In their area of nitrogen- and

sulfur-limited soils, starter fertilizer containing nitrogen and

sulfur substantially improved plant foliar cover and was

considered to be essential for successful seeding.

Figure 12.1—Relationship between turbidity in nephelometric turbidity units

(NTU) and suspended sediment parts per million (ppm) (Helvey and others

1985).

McCree Creek

Burns Creek

Fox Creek

Suspended sediment (mg/L)

Turbidity (NTU)

Turbidity < 10 NTU

Ÿ = –3.12 + 5.63 X

R2 = 0.723

Sy-x = 8.16

Turbidity ≥10 NTU

Ÿ = 4.57 + 4.64 x

R2 = 0.982

Sy-x = 70.5

5000

1000

500

100

50

10

5

1

1 5 10 50 100 500 1000

Fire Management
127

Table 12.3—Sediment yield after fire alone or in combination with other treatments

Pretreatment

Treatment Habitat Location or control Posttreatment Reference

 - - - - - - Pounds per acre per year - - - - - -

Prescribed Bluestem grasses Upper Coastal Plain, 74 – 84 1,542 – 5,759 Ursic 1969

burn northern Mississippi

Prescribed Post oak, hickories, Upper Coastal Plain, 148 – 300 868 – 1,179 Ursic 1970

underburn blackjack oak northern Mississippi

Prescribed fire, Juniper Central Texas Wright and others 1976

pile, and burn 3 to 4% slope 18 18

 8 to 20% slope 146 504

 37 to 61% slope 6 5,554

Wildfire Ponderosa pine Northwestern 3 1,254 Campbell and others 1977

 Arizona

Ponderosa pine, Eastern 12 – 35 146 – 2,100 Helvey 1980, Helvey and

 Douglas-fir Washington others 1985

Prescribed Loblolly pine Upper Piedmont, 19 20 Douglass and Van Lear 1983

underburn plantation South Carolina

Wildfire Chaparral Southern 12,500 98,160 Wells 1986

 California

Clearcut: broad- After clearcut: Southern Oregon 80 a 55 Amaranthus 1989

cast burned, vegetation: tanoak,

planted to madrone, chinkapin,

Douglas-fir, black oak, poison oak

wildfire

Adjacent forest: Douglas-fir 79 40 Amaranthus 1989

wildfire overstory; tanoak,

madrone, black oak

understory

Clearcut and Oak spp., Northwestern Not known Robichaud and Waldrop 1994

prescribed fire: shortleaf pine South Carolina

 Low severity 12.1

 High severity 502

a From October 13 to May 4, after September wildfire.

Chapter 12
128

Figure 12.2—Possible pathways of plant- and litter-contained nutrients

in response to combustion (Tiedemann 1981).

Temperature

Issues and Risks

Increases in streamwater temperature have important effects

on aquatic habitat and stream and lake eutrophication.

Eutrophication can adversely affect the color, taste, and

smell of drinking water. See chapter 2 for temperature

impacts on drinking water.

Findings from Studies

When riparian vegetation is removed by fire or other means,

the stream surface is exposed to direct solar radiation, and

stream temperatures increase (Levno and Rothacher 1969,

Swift and Messer 1971). For example, clearcutting and slash

burning increased stream temperatures by 13.0, 14.0, and

12.1 o

F (7.2, 7.7, and 6.7 oC) in June, July, and August, with

temperatures reaching a maximum of 75 oF (23.9 oC) in July

(Levno and Rothacher 1969). Helvey (1972) found that

during the first year after wildfire in eastern Washington,

stream temperature increased 10 oF (5.6 oC). In southern

Oregon, Amaranthus and others (1989) determined that

temperatures increased 6, 11, and 18 oF (3.3, 6.1, and 10 oC),

from a low temperature of 55 oF (12.8 oC) to a high temperature of 73 o

F (22.8 oC) after a wildfire. These temperature

changes have the potential to increase the rate of eutrophication if phosphate is present in abundance.

Chemical Water Quality

Several chemical constituents are likely to come from forest

and rangeland burning. The primary ones of concern are

nitrate (NO3

-1) and nitrite (NO2-1). Sulfate, pH, total

dissolved solids, chloride, iron, turbidity (discussed previously), and several other constituents can also be affected, as

can color, taste, and smell (see chapter 2). Phosphate (P) can

affect water quality because of its ability to affect color,

taste, and smell by accelerating the eutrophication process.

To understand the influence of fire on water quality, it is

important to understand some of the changes in plant, forest

floor, and soil nutrients during and after the combustion

process. Burning oxidizes organic material, resulting in

direct loss of elements to the atmosphere as volatilized

compounds above critical temperatures, as particulates are

carried away in smoke, or elements are converted to oxides

to the ash layer (DeBano and others 1998, Raison and others

1985, Tiedemann 1981) (fig. 12.2). Nitrogen, sulfur, and

potassium are all susceptible to volatilization loss by

burning (DeBano and others 1998, Raison and others 1985,

Tiedemann 1987). Nitrogen is lost when temperatures reach

400 o

F (204 oC) (DeBano and others 1998). At temperatures

as low as 700 oF (371 oC), loss of sulfur can be substantial

(Tiedemann 1987). As temperatures approach 1,475 oF

(802 oC), virtually all nitrogen and sulfur are volatilized. At

1,430 oF (776 oC), phosphorus and potassium are volatilized.

In ashes, relatively insoluble oxides of metallic cations, such

as calcium, potassium, magnesium, and iron, react with

water and carbon dioxide in the atmosphere and become

more soluble (DeBano and others 1998, Tiedemann 1981)

(fig. 12.2). This conversion increases potential for leaching

loss of nutrients from the ash into and through the soil

(DeBano and others 1998, Tiedemann 1981, Wells and

others 1979). Nutrients in the ash are also susceptible to loss

by surface erosion (Beschta 1990, DeBano and others 1998,

Tiedemann 1981, Tiedemann and others 1979, Wells and

others 1979).

The potential for increased nitrate in streamflow occurs

mainly because of accelerated mineralization and nitrification in soils after burning (Covington and Sackett 1986,

1992; DeBano and others 1998; Vitousek and Melillo 1979),

as well as reduced plant demand (Vitousek and Melillo

1979). This effect is short-lived, usually lasting only a year

or so (Monleon and others 1997).

Transport of nutrients to streams occurs both during and

after a wildland fire. Spencer and Hauer (1991) reported that

the source of nitrogen in streamwater during a fire appears

to be diffusion of smoke and gasses directly into the

streamwater, and that the source of phosphorus in streamwater appears to be from the leaching of ash deposited

directly into the stream. After a fire, nutrients from ash

deposition move from the soil into streamwater when

precipitation is adequate for percolation below the root

Soil and soil nutrients

Nutrients in

plants and litter

Combustion

Volatilization (N, P, S, K)

Deposition as

ash (Ca, Mg)

Surface

Leaching loss

Runoff and

erosion loss

Leaf

Uptake

fall

Fire Management
129

zone, and when the capacity of vegetation for uptake or soil

nutrient storage capacity, or both, are insufficient to retain

mobile nutrients carried into the soil (Beschta 1990, DeBano

and others 1998, Tiedemann and others 1979).

Issues and Risks

The issue is whether forest or rangeland fires degrade the

quality of source water for public consumption by the

introduction of additional chemical constituents. The risk is

when these additional chemical constituents—from a fire or

from fertilizer applied to establish vegetation in the burn

area—are combined with chemical constituents already

present, the source water supply may be degraded.

Findings from Studies

Immediately after a fire, the pH of streams may be affected

by direct ash deposition. In the first year after fire, increased

pH of the soil (Wells and others 1979) may also contribute

to increased streamwater pH. In all the studies we evaluated

(table 12.4), only one reported a notable increase in pH

values. During the first 8 months after the Entiat fires in

eastern Washington, Tiedemann (1973) detected transient

pH values up to 9.5. Two days after fertilization, they

detected a transient pH value of 9.2. In most studies pH

values were little changed by fire and fire-associated events.

Nitrogen

The forms of nitrogen that are of concern in drinking water

after fire are nitrate and nitrite. Values for nitrate generally

increased after fire but not to a level of concern (table 12.5),

except in nitrogen-saturated areas (see chapter 3). Stream

nitrate responses to prescribed fire are generally lower than

for wildfire. In an undisturbed ponderosa pine and Gambel

oak or both (P. ponderosa Dougl. ex Laws. and Q. gambelii

Nutt. or both) watershed in Arizona, Gottfried and DeBano

(1990) found that a fire resulted in only slight, but significant, increases in nitrate (table 12.5). Measures to protect

streams and riparian areas during prescribed burns with

unburned buffers could minimize effects of fire on stream

chemistry.

The most striking response of nitrate concentration in

streamflow after wildfire (table 12.5) was observed in

southern California (Riggan and others 1994). Moderate

burning resulted in a maximum nitrate concentration of 9.5

ppm, while severe burning resulted in a maximum concentration of 15.3 ppm in streamflow, compared to 2.5 ppm in

streamflow from an unburned control watershed. The

concentration of 15.3 ppm is above maximum contaminant

level for drinking water of 10 ppm (chapter 2, table 2.3).

Chronic atmospheric deposition of nitrogen pollutants on

these watersheds, which are east of Los Angeles, CA, have

caused their soils to become nitrogen saturated. Beschta

(1990) reached the same conclusion in his assessment of

streamflow nitrate responses to fire and associated treatments. Fenn and others (1998) have discussed excess

nitrogen in ecosystems in North America. These excess

levels can lead to leaching of nitrate, which ultimately can

find its way into streamwater (see chapter 3).

Fertilization after fire resulted in higher concentrations of

nitrate than fire alone (table 12.5) (Tiedemann 1973,

Tiedemann and others 1978). Nonetheless, Tiedemann

(1973) concluded that neither fire nor nitrogen fertilization

at levels less than 54 pounds per acre (60.5 kg per hectare)

of elemental nitrogen would probably have adverse effects

on nitrate concentrations in drinking water. Their research

was done in an area with nitrogen-limiting soils. In areas

experiencing nitrogen saturation, nitrogen fertilization may

aggravate nitrate levels in water and is not likely to stimulate revegetation.

Nitrite was reported by itself, rather than in combination

with nitrate, in only two studies that we found. At concentrations >1 ppm, nitrite can lead to serious illness in infants

(chapter 2, table 2.3). At the Lexington Reservoir, Santa

Clara County, CA, Taylor and others (1993) found nitrite

levels of 0.03 ppm after the watershed above the reservoir

was burned, while control levels were 0.01 ppm. Tiedemann

(1973) reported that nitrite concentrations were below the

levels of detection. The concentrations found do not appear

to be a concern.

Fire retardants containing nitrogen have the potential to

affect the quality of drinking water, but research on the

application of retardants to streams has focused on the

effects on fish and aquatic habitat (Buhl and Hamilton 1998;

Gaikowski and others 1996; McDonald and others 1996,

1997; Norris and Webb 1989; Norris and others 1978).

Several in vitro research projects evaluated the toxicity to

stream organisms of some retardant formulations. The tested

compounds were nonfoam retardants containing sulfate,

phosphate, and ammonium compounds; a retardant containing ammonium and phosphate compounds; and two foam

suppressant compounds (Buhl and Hamilton 1998;

Gaikowski and others 1996; McDonald and others 1996,

1997). Concentrations of nitrate rose from 0.08 to 3.93 ppm

after adding the nonfoam retardants. In addition, they found

Chapter 12
130

Table 12.4—The pH in water after fire alone or in combination with other treatments usually remains fairly

constant

Pretreatment

Treatment Habitat Location or control Posttreatment Reference

 - - - - - - - - - pH - - - - - - - - -

Wildfire Ponderosa pine, Eastern None given 7.2 – 8.5 Tiedemann 1973

 Wildfire and N Douglas-fir Washington

 fertilization None given 7.1 – 9.5 a

Wildfire and N Mixed conifer, Central Sierra ~7.0 – 6.2b ~7.0 – 6.6b Hoffman and Ferreira 1976

 fertilization shrub Nevada Mountains,

 California

Pile, burn Juniper Central Texas

 3 to 4% slope 7.3 7.3 Wright and others 1976

 8 to 20% slope 7.6 7.7

 37 to 61% slope 7.4 7.7

Wildfire Pine, spruce, Northeastern 6.2 6.1 – 6.3 Tarapchak and Wright 1977

 fir, aspen, birchc Minnesota lakes

Wildfire Ponderosa pine, Eastern 7.4 – 7.6 7.4 – 7.6 Tiedemann and others 1978

 Wildfire and N Douglas-fir Washington

 fertilization

Prescribed fire Ponderosa pine Central Arizona 6.2 6.4 Sims and others 1981

Pile, burn, Juniper Central Texas 7.1 7.3 Wright and others 1982

and seed

Clearcut, slash Western hemlock, Western British 6.8 7.8 Feller and Kimmins 1984

broadcast burned western red cedar, Columbia

 Douglas-fir

Yellowstone Subalpine lake Yellowstone Lake, 7.4 7.5 Lathrop 1994

 wildfires Yellowstone National

 Park, WY

~ = About or approximately.

aTransient pH value of 9.5 was observed second day after urea fertilization.

b From May to July during the summer following the August fire.

cCited in Wright and Watts 1969.

Fire Management
131

Table 12.5—Maximum nitrate-nitrogen concentration in water after fire alone or in combination with other treatments

Pretreatment

Treatment Habitat Location or control Posttreatment Reference

- - - - - Parts per million - - - - -

Clearcut, slash

burned Douglas-fir Western Oregon 0.1 0.43 Fredriksen 1971

Wildfire Ponderosa pine Eastern .016a .042 Tiedemann 1973

Wildfire and Douglas-fir Washington .005 .310a

 nitrogen

fertilization

Wildfire Mixed conifer, Central Sierra ~.6a b ~.12 Hoffman and Ferreira 1976

 shrub Nevada Mountains,

 California

Ponderosa pine Northwestern .086 .212 Campbell and others 1977

 Arizona

Pine, spruce, Northeastern .17 .08 – .17 Tarapchak and Wright 1977

 fir, aspen, birchc Minnesota lakes

Wildfire Ponderosa pine Eastern < .016a .56 Tiedemann and others 1978

Wildfire and Douglas-fir Washington < .016a .54 –1.47

nitrogen

fertilization

Prescribed fire Pine forest or Lower Coastal Plain, d .02 Richter and others 1982

 not given South Carolina

Prescribed Loblolly Upper Piedmont, .05 .05 Douglass and Van Lear 1983

underburn pine plantation South Carolina

Clearcut, slash Douglas-fir, Southern Idaho .02 .05 Clayton and Kennedy 1985

broadcast burned ponderosa pine

Prescribed burn, Ponderosa pine, Central Arizona 0.0013a 0.0029 Gottfried and DeBano 1990

moderate gambel oak

Wildfire Chaparral Lexington Reservoir, .02 .04 Taylor and others 1993

 Santa Clara County, CA

Prescribed White fir, giant Sequoia National .001 – .005 .010 – .394 Chorover and others 1994

underburn sequoia, red fir, Park, CA

 sugar pine,

 Jeffrey pine

Prescribed broadcast: Chaparral Southern Riggan and others 1994

Moderate burn California 2.5 9.5

Severe burn 2.5 15.3

Wildfire Lodgepole pine, Glacier National < .040 .124 – .312 Hauer and Spencer 1998

 Douglas-fir, Park, MT

 ponderosa pine,

 western larch

~ = About or approximately.

a Maximum level attained.

b

 Mean concentration from May to July after August fire.

c

 Cited in Wright and Watts (1969).

d Pretreatment not significantly different from posttreatment.

Chapter 12
132

Fire Management

in vitro nitrite reached concentrations as high as 33.2 ppm.

Accidental deposition of retardants in streams has produced

values of nitrate and ammonia sufficiently high to be of

concern in drinking water.2 Great caution needs to be

exerted to keep retardant chemicals out of streams that are

public drinking water sources.

Phosphorus

Phosphate, as a component of fire retardants, can lead to

eutrophication. See chapter 3 for discussion of phosphorus

impacts on drinking water. Prior to wildfire, phosphate

concentrations ranged from 0.007 ppm to 0.17 ppm

(Hoffman and Ferreira 1976, Tiedemann and others 1978,

Wright and others 1976). After wildfire, prescribed fire, or

clearcutting followed by broadcast burning, phosphate

concentrations stayed the same or increased only as high as

0.2 ppm (Longstreth and Patten 1975). Any phosphorus

added to the stream system may have been taken up by the

aquatic organisms and, therefore, little increase was detectable. We found no reports of changes in phosphate concentration as the result of an inadvertent application of retardant

directly into a stream.

Sulfur

The sulfate ion is relatively mobile in soil water systems

(Johnson and Cole 1977). Although not as well studied as

those for nitrogen, the mineralization processes for sulfur

are similar. In streamwater from wildland watersheds,

observed levels of sulfate are usually low (table 12.6).

Control or prefire values range from as low as 1.17 ppm to

as high as 66 ppm, while postfire values range from 1.7 ppm

to a high of 80.7 ppm, well below the recommended

secondary drinking water standard (250 ppm) (table 2.4).

Chloride

Chloride response to fire and clearcutting plus fire has been

documented in several studies, and all responses are low

(table 12.7). Chloride concentrations in control or prefire

samples ranged from 0.49 to 6.4 ppm, and the chloride

concentration in postfire samples ranged from 0.40 to

7.1 ppm (Lathrop 1994), well below the recommended

secondary drinking water standard (250 ppm) (table 2.4).

Lewis Lake in Yellowstone National Park, WY, with its

large volume of water, had the highest chloride values for

both the prefire and postfire periods among the data examined.

Total Dissolved Solids

Only two studies reported total dissolved solids; many other

studies measured some of the constituents of total dissolved

solids but not total dissolved solids per se. Hoffman and

Ferreira (1976) detected a total dissolved solids concentration of about 11 ppm in the control area and 13 ppm in the

burned area, which had been a mixed conifer and shrub

stand in Kings Canyon National Park, CA. Lathrop (1994)

found Yellowstone Lake in Yellowstone National Park and

Lewis Lake had pretreatment total dissolved solids concentrations of 65.8 and 70 ppm. The total dissolved solids

concentrations after the fires were 64.8 and 76 ppm, well

below the recommended secondary drinking water standard

(500 ppm) (table 2.4).

Trace Elements

Fredriksen’s (1971) results raise a question about how well

we understand the responses of micronutrients or trace

elements to fire or to fire after clearcutting. In his stream

chemistry profile after clearcutting and broadcast burning,

he documented a maximum concentration of manganese of

0.44 ppm, exceeding the recommended secondary drinking

water standard (0.05 ppm) (table 2.4), which may raise

palatability issues but is not a health risk. There are established drinking water standards for 14 additional trace

constituents, including heavy metals. Information on the

effects of these elements after a forest or rangeland fire on

drinking water quality is lacking.

Effects on Ground Water

Little research has been conducted on the effects of fire, fire

suppression, and fire rehabilitation activities on ground

water quality. It is reasonable to expect that fire will have

little effect on ground water quality. A possible, but unlikely,

scenario would be a fire followed by an intense longduration precipitation event sufficient to cause major

flooding, which could contaminate ground water. In such a

case, the fire sets the stage for contamination of the ground

water source.

2

 Labat-Anderson Incorporated. 1994. Chemicals used in wildland fire

suppression: a risk assessment. Prepared for: Fire and Aviation Management, U.S. Department of Agriculture, Forest Service. Contract

53-3187-9-30; Task 93-02. 187 p. Prepared by: Labat-Anderson Incorporated, 2200 Clarendon Boulevard, Suite 900, Arlington, VA 22202.
133

Table 12.7—Chloride concentration in water after fire alone

Pretreatment

Treatment Habitat Location or control Posttreatment Reference

- - - - - Parts per million - - - -

Wildfire Taiga Interior Alaska 0.9 – 5.0 1.2 – 4.6 Lotspeich and others 1970

Mixed conifer, Central Sierra .6 1.0 Hoffman and Ferreira 1976

 shrub Nevada Mountains,

 CA

Pine, spruce, fir, Northeastern .80 – .89 1.24 Tarapchak and Wright 1977

 aspen, bircha Minnesota lakes

Prescribed White fir, giant Sequoia National .49 – .56 .40 – 2.78 Chorover and others 1994

underburn sequoia, red fir, Park, CA

 sugar pine,

 Jeffrey pine

Yellowstone Subalpine lakes Yellowstone Lake 5.1 3.6 Lathrop 1994

wildfires Lewis Lake, 6.4 7.1

 Yellowstone National

 Park, WY

a Cited in Wright and Watts 1969.

Table 12.6—Sulfate concentration in water after fire alone

Pretreatment

Treatment Habitat Location or control Posttreatment Reference

- - - - - Parts per million - - - -

Wildfire Taiga Interior Alaska 7.12 – 66 8.3 –80.7 Lotspeich and others 1970

Mixed conifer, Central Sierra 1.5 1.7 Hoffman and Ferreira 1976

 shrub Nevada Mountains,

 CA

Pine, spruce, fir, Northeastern 1.17 1.79 – 1.86 Tarapchak and Wright 1977

 aspen, bircha Minnesota lakes

Prescribed White fir, giant Sequoia National .26 .37, .30, .45b Chorover and others 1994

underburn sequoia, red fir, Park, CA, Log Creek,

 sugar pine, control

 Jeffrey pine

White fir, fewer Tharp’s Creek, burn .24 9.68, 1.32, 2.15b

 Giant sequoia

Yellowstone Subalpine lakes Yellowstone Lake 8.9 6.4c Lathrop 1994

wildfires Lewis Lake, 4.0 3.0

 Yellowstone National

 Park, WY

a

 Cited in Wright and Watts 1969.

b Postburn years one, two, and three, in sequence.

c Average of reported median values from four areas of Yellowstone Lake.

Chapter 12
134

Fire Management

Reliability and Limitations of Findings

The results of research on the effects of fire on drinking

water quality are strong and consistent, especially from the

Pacific Northwest, the Rocky Mountains, and the Southwest. The results indicate that the effects of fire and fire

management practices on water quality are similar within

each of these three large areas. Data from the Southeast are

somewhat more limited in spite of the region’s extensive

prescribed fire program. In Alaska, water-quality research

after fire has been minimal, even though the area has many

wildfires.

With the changes in pH, nitrate, sulfate, and chloride so

consistently small, a land manager can safely assume that

effects will be similar to those found in the literature, if the

treatments and fire severity, slope, and soil and vegetation

types are comparable. In areas likely to be nitrogen saturated, such as areas of high soil concentrations of nitrogen

from chronic atmospheric deposition, nitrate concentrations

in streamwater after a fire may exceed the established

maximum contaminant level of 10 ppm (chapter 2, table

2.3). In areas of suspected nitrogen saturation, common

sense tells us that nitrogen-containing fertilizer should not

be applied. Application of nitrogen fertilizer would exacerbate the risk of degrading source water supplies. See chapter

3 for more discussion of nitrogen saturation.

Results of previous wildland fires can be used as a basis for

estimating the effects of new fires on drinking water quality.

Fires need to be of the same type; that is, previous wildfires

should be used as the comparison basis for new wildfires,

and previous prescribed fires as the basis for new prescribed

fires. The more factors, such as slope and vegetation, among

others, that match between the previously documented fires

and the new fires, the closer the approximation. Nevertheless, results and predictions based on limited data must be

used cautiously. In two studies (Beschta 1980, Helvey and

others 1985), the researchers specifically caution against

extrapolating results of turbidity and sediment research

beyond the watersheds in which the research was conducted.

Research Needs

1. Research methods need to be carefully selected for

measurements of sediment. Suspended sediment concentration is used in some studies while turbidity, the

standard measurement, is used in others. At this time,

regression relationships between sediment, in parts per

million, and turbidity, in NTU’s, need to be developed for

each individual watershed or stream system. Interpretation of future research results will be facilitated when all

measurements are reported in the standard NTU’s.

2. Areas with chronic atmospheric deposition, such as those

studied by Riggan and others (1994), need further

research into the relations between fire and nitrogen

release into streams.

3. We have little information on the abundance of trace

elements (micronutrients) after fire. When elements, such

as lead, copper, fluoride, manganese, iron, zinc, and

mercury, among others, are above certain levels, they are

important potential contaminants in drinking water

supplies. We do not understand the effects of fire in

combination with other treatments on micronutrients.

Effects may be particularly important for some of the

heavy-metal trace elements.

4. The inadvertent application of fire retardants directly into

a stream can produce increased levels of nitrate and

possibly sulfate, phosphate, and some trace elements.

Information is needed about the potential effects of

specific retardants on drinking water quality.

5. The BAER practices, particularly the use of contourfelled erosion barriers, need to be systematically studied

to determine their effectiveness for reducing storm

runoff, erosion, and sediment movement, which pose a

risk to the quality of source water for public water

supplies.

Key Points

1. When a wildland fire occurs, the principal concerns

for change in drinking water quality are: (1) the introduction of sediment; and (2) the potential introduction of

nitrates, especially if in areas with chronic atmospheric

deposition.

2. As we considered the above types of fire effects on

drinking water, several concepts important to the land

manager became apparent. The magnitude of the effects

of fire on water quality is primarily driven by fire

severity, and not necessarily by fire intensity. Fire

severity is a qualitative term describing the amount of

fuel consumed, while fire intensity is a quantitative

measure of the rate of heat release. In other words, the

more severe the fire the greater the amount of fuel

consumed and nutrients released and the more susceptible

the site is to erosion of soil and nutrients into the stream

where it could potentially affect water quality. Wildfires

usually are more severe than prescribed fires, and, as a

result, they are more likely to produce significant effects

on water quality. On the other hand, prescribed fires are

designed to be less severe and would be expected to

produce less effect on water quality. Use of prescribed

fire allows the manager the opportunity to control the
135

severity of the fire and to avoid creating large areas

burned at high severity. The degree of fire severity is also

related to the vegetation type. For example, in grasslands

the differences between prescribed fire and wildfire are

probably small. In forested environments, the magnitude

of the effects of fire on water quality will probably be

much lower after a prescribed fire than after a wildfire

because a larger amount of fuel may be consumed in a

wildfire. Canopy-consuming wildfires would be expected

to be of the most concern to managers because of the loss

of canopy coupled with the destruction of soil aggregates.

These losses present the worst-case scenario in terms of

water quality. The differences between wild and prescribed fire in shrublands are probably intermediate

between those seen in grass and forest environments.

3. Another important determinant of the magnitude of the

effects of fire on water quality is slope. Steepness of the

slope has a significant influence on movement of soil and

nutrients into stream channels where it can affect water

quality. Wright and others (1976) found that as slope

increased in a prescribed fire, erosion from slopes is

accelerated. If at all possible, the vegetative canopy on

steep, erodible slopes needs to be maintained, particularly

if adequate streamside buffer strips do not exist to trap

the large amounts of sediment and nutrients than can be

transported quickly into the stream channel. It is important to maintain streamside buffer strips whenever

possible, especially when developing prescribed fire

plans. These buffer strips will capture much of the

sediment and nutrients from burned upslope areas.

4. Two more concerns, which are more site-specific, deal

with soils. Both the general type of soil and a soil’s

propensity to develop water repellency can be determinants of the magnitude of the effects of fire on drinking

water quality. When sandy soils are burned, nutrient

transport and loss are rapid. These soils do not have the

ability to capture and hold nutrients, but, rather, allow the

nutrients to move into the ground water and eventually

into nearby streams. Additionally, in areas with sandy

soils, which contain few nutrients, most of the nutrient

capital is stored aboveground. A severe fire volatilizes

many of these nutrients, impoverishing the site, while

adding to the nutrient load in streams. Prescribed fires in

these areas need to be very carefully planned to retain as

many nutrients on site as possible through the use of lowseverity fires.

5. If a site is close to nitrogen saturation, it is possible to

exceed maximum contamination levels for drinking water

of nitrate (10 ppm) after a severe fire. Such areas should

not have nitrogen-containing fertilizer applied after the

fire. See chapter 3 for more discussion of nitrogen

saturation.

6. The propensity for a site to develop water repellency after

fire must be considered. Water-repellent soils do not

allow precipitation to penetrate down into the soil and,

therefore, are conducive to erosion. Such sites can put

large amounts of sediment and nutrients into surface

water.

7. Finally, heavy rain on recently burned land can seriously

degrade water quality. The effects of the Buffalo Creek

wildfire on the water supply for Denver, noted in the first

paragraph of this chapter, demonstrated these effects on a

water supply. Severe erosion and runoff are not limited to

wildfire sites alone. If the storm delivers large amounts of

precipitation or is sufficiently intense, accelerated erosion

and runoff can occur after a carefully planned prescribed

fire. Conversely, if below-average precipitation occurs

after a wildfire, there may not be a substantial increase in

erosion and runoff.

8. The land manager can influence the effects of fire on

drinking water quality by careful prescribed burning.

Limiting fire severity, avoiding burning on steep slopes,

and limiting burning on sandy or potentially waterrepellent soils will reduce the magnitude of the effects of

fire on water quality.

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Fire Management
139

Chapter 13

 Pesticides

J.L. Michael1

Introduction

On forest and grassland, management often must protect

desirable vegetation from pathogens, competing vegetation,

insects, and animals. Vegetation also is managed to clear

road and utility rights-of-way, to improve recreation areas

and wildlife habitat, and to control noxious weeds. Pesticides offer inexpensive and effective ways of getting these

jobs done. The Forest Service requires: (1) training of

personnel who recommend and use pesticides, (2) applicator

certification, and (3) safety plans to assure the safety of

personnel and the protection of environmental values like

drinking water quality. Nonnational forest land is treated

with pesticides for many of the same purposes, but often

more intensively.

The Federal Insecticide, Fungicide, and Rodenticide Act, as

amended (FIFRA) (Public Law 92–516, and 40 CFR 158)

allows the registration of pesticides for use in the United

States. The registration process is an extraordinary one that

requires years of testing before sufficient efficacy, environmental safety, toxicology, and public safety data can be

collected and evaluated in the support of registration of a

new pesticide. While this process is designed to assure

safety, new and old pesticides, following registration,

continue to be studied by researchers in private, State, and

Federal agencies in an effort to identify any potential

environmental or toxicological problems. An integral part of

protecting public health and environmental values during

pesticide use is the requirement that they must be applied

according to directions approved by the U.S. Environmental

Protection Agency (EPA) and included on the label of every

registered pesticide. Under FIFRA, pesticide labels are

legally binding documents, and any infraction of the

directions for application is a violation of law. Users of

pesticides must exercise extreme caution in following label

directions and must also exercise good judgement, especially when pesticide use is planned in an area near municipal water supplies. In addition, pesticide users must provide

adequate handling facilities for mixing and storage and be

well prepared to deal with spills.

To meet the minimum requirements of FIFRA at the State

level, the EPA has established and maintains cooperative

enforcement agreements for pesticide use inspections,

producer establishment inspections, marketplace surveillance, applicator certification, and experimental use inspections. State government is responsible for (1) certification of

pesticide applicators, (2) enforcement of FIFRA pesticide

use regulations and inspections, (3) endangered species

considerations, (4) worker protection, and (5) ground water

protection.

When forestry pesticides are used near water on Federal,

State, or privately owned land, buffer zones are left between

the treated areas and the water resource (see chapter 5). The

width of the buffer varies with site conditions, site sensitivity, and local or State recommendations. National forests in

some States use more conservative buffers than those

recommended by the State. Comerford and others (1992)

have reviewed many agricultural studies in an attempt to

draw inferences regarding effectiveness of buffer strips in

mediating stream contamination. However, relatively little

research data are available on effectiveness of buffers on

forest sites. It, therefore, is not possible to determine the

minimum buffer width to protect streams from either

pesticide or sediment contamination.

Issues and Risks, Pesticide Application

Approximately 16 percent of the 3.6 million square miles of

land in the United States is treated with pesticides annually

(Pimentel and Levitan 1986). The most intensive use of

pesticides occurs on land occupied by households. Household tracts account for only 0.4 percent of all land but

receive 12 percent of all pesticides used in the United States.

Agricultural land (52 percent of all land) is the next most

intensively treated, receiving 75 percent of all pesticides

used. Government and industrial land (16 percent of all

land) receives 12 percent of all pesticides. The least intensive use of pesticides occurs on forest land (32 percent of

the land). Pimentel and Levitan (1986) point out that forest

land receives only 1 percent of all pesticides used in the

United States and that <1 percent of all forest land is treated

annually. 1

 Research Ecologist, USDA Forest Service, Southern Research Station,

Auburn, AL.
140

A wide variety of pesticides are used on forest and grasslands. Table 13.1 lists these pesticides and the purposes

for which they are used. Maximum Contaminant Levels

(MCL’s), established by EPA, are listed in chapter 2,

table 2.3.

Biological control agents, including Bacillus thuringiensis

(Bt) and nucleopolyhedrosis virus (Npv), are used for

control of western spruce budworms and gypsy moths.

While these are pathogens of insects, they have no known

impacts on drinking water quality.

Plant pathogens represent a potential problem throughout

forest and rangeland ecosystems, but their destructive

impacts are most severe in seedling nurseries and seed

storage facilities. Fungicides and fumigants are used to

control these pathogens on seeds, in seedling nurseries, in

greenhouses, in seed tree nurseries, and on individual trees.

In 1997, the Forest Service treated 35 greenhouses with

fungicides and fumigants including benomyl, chlorothalonil,

dicloran, iprodione, metalaxyl, propiconazole, thiophanate

methyl, and triadimefon. Most fungicide and fumigant use

occurs on small acreages in nurseries for disease control.

Small amounts of strychnine and putrescent egg solids were

used over extensive acreages for animal damage control.

While very small amounts of strychnine were used over vast

acreages, it is very toxic (table 13.2). Putrescent egg solids,

by comparison, are derived from food products and the EPA

has waived toxicology requirements. These two products

accounted for more than 96 percent of the active ingredients

used in protection of vegetation from animals. Insect control

relied mainly on biological agents, but some insecticides

and oils were used. The insecticides included carbaryl and

chlorpyrifos. Dormant oil was used for control of a variety

of insects and their eggs. These three (carbaryl, chlorpyrifos,

and dormant oil) represent 94.6 percent of all chemical

insecticides used on National Forest System (NFS) land.

Vegetation management is frequently taken to mean the

control of competing vegetation in timber management

programs. On NFS land, more than three times as much land

was treated for protection of vegetation from animals and

insects and to control noxious weeds, than for control of

competing vegetation in timber management programs in

fiscal year 1997 (U.S. Department of Agriculture, Forest

Service 1998). Competing vegetation can be controlled with

herbicides, algicides, and plant growth regulators. Table

13.1 shows management objectives for herbicide use on

NFS land in fiscal year 1997. Many acres are treated for

timber management, principally planting site preparation

and release of crop trees. Such treatments usually occur only

once or twice over a rotation. Rotation length depends on

tree species, site productivity, and management objectives.

The rotation may be as short as 20, or longer than 150 years.

Thus, herbicides are used in timber management only once

or twice in 20 or more years. Treatments were for site

preparation, conifer release, and hardwood release.

Noxious weed control is often accomplished by treatment

with herbicides. Noxious weeds are usually nonnative plants

that, lacking natural controls, spread quickly and take over

or ruin habitat for native plants. They generally possess one

or more of the following characteristics: aggressive and

difficult to manage, poisonous, toxic, parasitic, and a carrier

or host of serious insects or disease. There are 74 terrestrial

species on the Federal noxious weed list, including kudzu.

The frequency of noxious-weed treatment varies by species.

In southern forests, kudzu requires annual treatment over

several years for effective control. Typically, attempts to

control noxious weeds do not eradicate them, but bring them

under enough control to reduce immediate problems. Timber

management (45.2 percent) and noxious weed control (48.8

percent) accounted for 94 percent of all acres treated with

herbicides.

Protecting forests and seedlings from animal pests is the

single largest component of the vegetation management

program. Rabbits and deer were the most common target

mammals, while western spruce budworms and gypsy moths

were the principal target insects.

One major issue with pesticide use is the impact on drinking

water quality. To adversely impact drinking water, pesticides

must (1) be harmful to humans, and (2) reach drinking water

at concentrations exceeding toxic levels for humans.

Issues and Risks, Toxicity

The toxicity of a chemical is a measure of its ability to harm

individuals of the species under consideration. This harm

may come from interference with biochemical processes,

interruption of enzyme function, or organ damage. Toxicity

may be expressed in many ways. Probably the best known

term is LD50, the dose at which 50 percent of the test

animals are killed. More useful terms have come into

popular usage in the last decade: no observed effect level

(NOEL), no observed adverse effect level (NOAEL), lowest

observed adverse effect level (LOAEL), reference dose

(RfD), and, relating specifically to water, the health advisory

level (HA or HAL). The EPA uses these terms extensively in

risk assessment programs to indicate levels of exposure

deemed safe for humans, including sensitive individuals.

They are derived from toxicological test data and have builtin safety factors ranging upward from 10, depending on

EPA’s evaluation of the reliability of the test data.

Pesticides
141

The NOEL is determined from animal studies in which a

range of doses is given daily; some doses cause adverse

effects and others do not (U.S. EPA 1993). The NOAEL is

derived from the test data where all doses have some effect,

but some of the observed effects are not considered adverse

to health. When EPA has data from a number of these tests,

the lowest NOEL or NOAEL is divided by a safety factor of

at least 100 to determine the RfD. The RfD is an estimate of

a daily exposure to humans that is likely to be without an

appreciable risk of deleterious effects during a lifetime.

Drinking water standards are calculated for humans by

assuming that an adult weighs 155 pounds and consumes

2 pints of water per day, and a child weighs 22 pounds and

consumes 1 pint of water per day over the period of exposure. The HAL’s are calculated for 1 day, 10 days, longer

term (10 percent of life expectancy), or lifetimes (70 years)

by dividing the NOAEL or LOAEL by a safety factor and

multiplying the resulting value by the ratio of body weight

to amount of water consumed daily (U.S. EPA 1993). The

safety factor can range from as low as 1, but is rarely <10,

and goes as high as 10,000, depending on the available

Table 13.1—Management uses of pesticides commonly used on national forests

Pesticide Vegetation management use(s)

2, 4-D Housekeeping and facilities maintenance, noxious weed control, nursery weed control,

 recreation improvement, right-of-way vegetation management, seed orchard protection,

 agricultural weed control, other vegetation management

Bacillus thuringiensis Insect suppression

Borax Disease control

Carbaryl Insect suppression in the field and in greenhouses, nursery insect control

Chloropicrin Nursery disease control

Chlorpyrifos Housekeeping and facilities maintenance, insect control, nursery insect control

Clopyralid Housekeeping and facilities maintenance, noxious weed control, nursery weed control,

 right-of-way vegetation management, wildlife habitat improvement

Dazomet Fungus control, nursery disease control, soil fumigation

Dicamba Noxious weed control, other vegetation management

Dormant oil Insect control

Hexazinone Wildlife habitat improvement, site preparation, conifer release

Imazapyr Conifer release, hardwood release, hardwood control, noxious weed control, site

 preparation

Methyl bromide Nursery disease control, soil fumigation

Metsulfuron Noxious weed control

Nucleopolyhedrosis virus Insect suppression

Picloram Noxious weed control, right-of-way vegetation management, weed control, wildlife

 habitat improvement

Putrescent egg solids Animal damage control

Strychnine Animal damage control, seed orchard protection

Thiram Animal damage control, fungus control, nursery disease control

Triclopyr Conifer release, hardwood control, hardwood release, noxious weed control, recreation

 improvement, right-of-way vegetation management, seed orchard protection, site

 preparation, thinning, general weed control, wildlife habitat improvement

Zinc phosphide Animal damage control

Chapter 13
142

toxicological data. A safety factor of 10 is used when good

NOAEL data are based on human exposures and are

supported by chronic or subchronic data in other species.

When NOAEL’s are available for one or more animal

species but not humans and good data for LOAEL in

humans is available, a safety factor of 100 is used. When

good chronic data are available identifying an LOAEL but

not an NOAEL for one or more animal species, a safety

factor of 1,000 is used. For situations where good chronic

data are absent, but subchronic data identify an LOAEL but

not an NOAEL, the safety factor of 10,000 is used. The

EPA’s estimates of safe levels for daily exposure to the

pesticides most widely used on NFS land are summarized in

table 13.2. Of the pesticides listed in table 13.2, elemental

boron (potentially from borax) and methyl bromide are

listed in EPA’s drinking water contaminant candidate list for

consideration for possible regulation. The MCL’s have been

established for 2,4-D [0.070 milligrams (mg) per liter],

glyphosate (0.700 mg per liter), and picloram (0.500 mg per

liter) and these are the same as the already established

lifetime HAL’s (table 13.2). Information on specific

pesticides can be retrieved from the National Pesticides

Telecommunication Network at http://ace.orst.edu/info/nptn,

EPA site at http://www.epa.gov/epahome/search.html,

Extension Toxicology Network at http://ace.orst.edu/info/

extoxnet, Material Data Safety Sheets at http://siri.uvm.edu/

msds, USDA Forest Service at http://www.fs.fed.us/

foresthealth/pesticide, and many others.

Table 13.2—Estimates of safe levels for daily exposure to the 20 pesticides most used on National Forest

System lands in fiscal year 1997 in the vegetation management program

Lifetime

Pesticide RfD NOEL NOAEL HAL Reference

- - - - - Milligrams per kilogram - - - - - mg/L

Borax 0.09 NA 8.8 0.60a U.S. EPA 1990

Carbaryl .1 NA 9.6 .700 U.S. EPA 1989

Chloropicrinb NA NA NA NA

Clopyralid NA NA NA NA

Chlorpyrifos .003 0.03 NA .020 U.S. EPA 1993

2,4-Dc .01 NA 1 .070 U.S. EPA 1989

Dazometb NA NA NA NA

Dicamba .03 NA 3 .200 U.S. EPA 1989

Dormant oil NA NA NA NA

Glyphosatec .1 20 NA .700 U.S. EPA 1989

Hexazinone .05 5 NA .400 U.S. EPA 1996

Imazapyr NA 250 NA NA U.S. EPA 1997

Methyl bromide .0014 NA 1.4 .010 U.S. EPA 1990

Metsulfuron .25 25 NA NA U.S. EPA 1988b

Picloramc .007 7 NA .500 U.S. EPA 1988a

Putrescent egg

solids NAd NA NA NA

Strychnine .0003 None None NA U.S. EPA 1998a

Thiram .005 5 NA NA U.S. EPA 1992

Triclopyr .05 5 NA NA U.S. EPA 1998b

Zinc phosphide .0003 None None NA U.S. EPA 1998a

HAL = health advisory level; NA = not available; NOAEL = no observed adverse effect level; NOEL = no observed effect level;

RfD = reference dose.

a HAL for elemental boron.

b These fumigants are not expected to get into water.

c Maximum contaminant levels for glyphosate (0.700 mg per liter), 2,4-D (0.070 mg per liter), and picloram (0.500 mg per liter) are discussed

in chapter 2.

d Made from food products; toxicology was waived by U.S. Environmental Protection Agency.

Pesticides
143

Findings from Studies

Pesticides used by the NFS in vegetation management are

used around the World in agricultural, forest, range, and

urban applications. Some have been found in surface water,

shallow ground water, and even in shallow wells (<30 ft),

but in concentrations far below levels harmful to human

health, and the occurrence is infrequent. Table 13.3 summarizes reports of pesticides from table 13.2 that have been

detected in water in the United States.

Larson and others (1997) summarized the results of 236

studies throughout the United States on pesticide contamination of surface water by listing the maximum observed

concentrations from each study. These studies were located

principally around large river drainage basins and, therefore,

represent cumulative pesticide contributions from a wide

variety of uses. Monitoring results were reported for 52

pesticides approved for agricultural, urban, and forestry use

and their metabolic byproducts. Of the pesticides listed in

table 13.2, only six were reported to be present in surface

water by Larson and others (1997). They were carbaryl,

1 report; hexazinone, 1 report; chlorpyrifos, 3 reports;

picloram, 4 reports; dicamba, 5 reports; and 2,4-D, 24

reports. None of the reported concentrations exceeded EPA

safe levels for human health except where application

included placement directly in stream channels and most

were <0.002 mg per liter. It is important to recognize that

surface water is not necessarily drinking water. The studies

summarized by Larson and others (1997) dealt with surface

water, principally in lakes, reservoirs, and rivers, which

would be treated prior to use for drinking. Thus, use of these

pesticides according to label directions has not resulted in

impairment of drinking water.

Reports of pesticide contamination of water are usually from

agricultural (Kolpin and others 1997, Koterba and others

1993) or urban applications (Bruce and McMahon 1996),

but the potential exists for contamination from forest

vegetation management. Water from forests is generally

much less contaminated than water from other land uses.

Several studies on forest sites listed in table 13.3 present

data for water collected directly from treated areas. The

concentration of pesticides can appear quite high compared

to samples taken from large rivers and lakes. Pesticide

concentrations are greatly reduced by dilution as they move

from the treated sites to downstream locations. Degradation

of pesticides by biological, hydrolytic, and photolytic routes

also contributes to downstream reductions in pesticide

concentrations.

From 1985–87, Cavalier and others (1989) monitored 119

wells, springs, and municipal water supplies for occurrence

of pesticides in drinking water throughout the State of

Arkansas. Monitored wells were generally located in the

eastern portion of Arkansas, but eight wells were located in

the Ouachita National Forest. Only sites considered highly

susceptible to contamination from pesticide use were

monitored, and these included domestic, municipal, and

irrigation wells. Detection limits for the three forestry

pesticides monitored (2,4-D, hexazinone, and picloram)

ranged from 70 to 800 times lower than their HAL’s. They

did not detect well water contamination from any of the 18

pesticides for which they monitored. Failure to detect

pesticides in these wells believed to be at high risk for

contamination is a very strong indicator that ground water is

not at risk from forestry pesticides used according to label

directions.

Michael and Neary (1993) reported on 23 studies conducted

on industrial forests in the South in which whole watersheds

received herbicide treatment. Water flowing from the sites

was sampled near the downstream edge of the treatments.

The watersheds were relatively small (<300 acres) and the

ephemeral to first-order streams draining these watersheds

were too small to be public drinking water sources, but their

flow reached downstream reservoirs. The maximum

observed hexazinone, imazapyr, picloram, and sulfometuron

concentrations in streams on these treated sites did not

exceed HAL’s, except for one case in which hexazinone was

experimentally applied directly to the stream channel. Even

in this case in which hexazinone was applied directly to the

stream at a very high rate, drinking water standards were

exceeded for only a few hours. In another study, picloram

was accidentally applied directly to streams, but maximum

picloram concentrations did not exceed HAL’s during the

year after application.

Bush and others (1990) reported on use of hexazinone on

two Coastal Plain sites (deep sand and sandy loam soils) that

were monitored for impacts on ground water. Hexazinone

was not detected in ground water at the South Carolina site

for 2 years after application. In Florida, hexazinone was

found infrequently in shallow test wells at concentrations up

to 0.035 mg per liter, much lower than the safe levels for

daily exposure (0.400 mg per liter). Water from these sites

drains into other creeks and rivers and is diluted before

entering reservoirs.

Michael and others (1999) reported the dilution of

hexazinone downstream of treated sites. One mile below

the treated site, hexazinone concentrations were diluted to

one-third to one-fifth the concentration observed on the

treated site. Hexazinone was applied for site preparation at

6 pounds active ingredient (ai) per acre to clay loam soils,

a rate three times the normal, and it was applied directly to

Chapter 13
144

Table 13.3—Frequency and occurrence of surface and ground water contamination from pesticide use in North

Americaa

Water

Pesticide type Location Maximum Range Comments Reference

- - - - - Milligrams per liter - - - - -

2, 4-D S Large river basins 0.0075 0.00004 – 0.0075 Twenty-four reports of mainly Larson and others 1997

throughout the urban, suburban, agricultural

United States sources

S Streams in Oregon 2.0 ND – 2.0 Highest concentrations observed Norris 1981

and California from forest areas where no

attempt was made to prevent

application to water

G Saskatchewan,

Canada .0000007 NG Natural spring flow Wood and Anthony 1997

G Connecticut, Iowa, .049 .0002 – .049 Well water samples, except for Funari and others 1995

Kansas, Maine, South Dakota, from shallow

Mississippi, South sand and gravel aquifer

Dakota

Borax NR NR NR NR NR

Carbaryl S Mississippi River .0001 NG One report Larson and others 1997

S New Brunswick, .314 NG Aerial spray spruce budworm Sundaram and Szeto 1987

 Canada control

S New Brunswick,

Canada .314 .123 – .314 Budworm control Holmes and others 1981

Chloropicrin NR NR NR NR NR

Chlorpyrifos S Mississippi River, .00015 .00004 – .00015 Three reports Larson and others 1997

the lower Colorado

River, rivers and lakes

in Kansas, irrigation

ditches in California

Arizona, Nevada

Clopyralid NR NR NR NR NR

Dazomet NR NR NR NR NR

Dicamba S USFS land near .037 .006 – .037 Treated 166 ac of 603-ac forest Norris 1975

Hebo, OR catchment; highest concentration

diluted to 0.006 mg/L 2.2 mi

downstream

Glyphosate S 45-ha coastal British .162 .0032 – .162 Highest concentration in streams Feng and others 1990

Columbian catchment intentionally sprayed, lowest

in streams with smz

S Quebec, Canada 3.080 .078 – 3.08 Nine of 36 streams contained Leveille and others 1993

glyphosate after forest spraying

S Ohio 5.2 NG No-tillage establishment of fescue Edwards and others 1980

S Georgia .035 NG Forest sites for scrub-hardwood Newton and others 1994

Michigan 1.237 NG control and direct spray of

Oregon .031 NG streams

continued

Pesticides
145

Table 13.3—Frequency and occurrence of surface and ground water contamination from pesticide use in North

Americaa (continued)

Water

Pesticide type Location Maximum Range Comments Reference

- - - - - Milligrams per liter - - - - -

Glyphosate

 (cont.) G Newfoundland, 0.045 0.004 – 0.045 Application of 4 lbs ai/ac to power Smith and others 1996

Canada substations resulted in contamination of water in monitoring wells

Hexazinone S Mississippi River .00007 NG Detected in five tributaries Larson and others 1997

S Alabama, Florida, .037 .0013 – 0.037 Seven reports, each treated Michael and Neary 1993

Georgia catchment containing ephemeral/

first-order streams

S Alabama 2.400 NG Applied directly to ephemeral Miller and Bace 1980

channel and in first runoff water

S Alabama .473 .422 – .473 Ephemeral/first-order stream in Michael and others 1999

catchments treated with 3x rate of

hexazinone in liquid and pellet

formulation with accidental

application to streams

S Arkansas .014 NG 11.5-ha watershed drained by Bouchard and others 1985

ephemeral to first-order stream

S Georgia .442 NG Ephemeral/first-order stream Neary and others 1986

in treated catchment, pellets

applied to stream channel

G NGb .009 NG Only one value reported from a Funari and others 1995

single study

Imazapyr S Alabama .680 .130 – .680 Two reports, each treated catch- Michael and Neary 1993

ment containing ephemeral/

first-order streams, herbicide

accidentally applied to stream

channel

Methyl

bromide NR NR NR NR NR

Metsulfuron S Central Florida .008 NG Water in surface depression Michael and others 1991

G .002 NG in slash pine site and 1 of 207

shallow (6-ft) well samples

Picloram S North-central

Arizona .32 NG Pinyon-juniper site Johnsen 1980

S Streams and rivers .005 .00001 – .005 Four reports from mainly range- Larson and others 1997

in North Dakota, land uses

Wyoming, Montana

S Alabama .442 NG Pellets accidentally applied Michael and others 1989

directly to forest stream

S Georgia, Kentucky, .021 ND – .021 Six study catchments with Michael and Neary 1993

Tennessee ephemeral/first-order stream in

each treated forest catchment

S North Carolina .01 NG Ephemeral/first-order stream in Neary and others 1985

treated forest catchment

continued

Chapter 13
146

a stream segment, resulting in a maximum observed on-site

concentration of 0.473 mg per liter. This was slightly more

than the lifetime HAL but considerably below the longer

term HAL of 9.0 mg per liter (U.S. EPA 1990). Following

the application, on-site stream concentrations did not exceed

the lifetime HAL.

Norris (1975) reported contamination of streamflow with

dicamba used for control of hardwoods on silty clay loam

soils in Oregon. On a 603-acre watershed, 166 acres were

aerially sprayed with 1 pound ai per acre [1.1 kilograms (kg)

per hectare] of dicamba. A small stream segment was also

sprayed causing detectable dicamba residues 2 hours after

application began, approximately 0.8 miles (1.3 kilometers)

downstream. Concentrations rose for approximately 5.2

hours after treatment began and reached a maximum

concentration of 0.037 mg per liter, less than one-fifth of

the HAL (0.200 mg per liter). No dicamba residues were

detected beyond 11 days after treatment.

Glyphosate and 2,4-D have aquatic labels, which permit

direct application to water. Stanley and others (1974) found

that when 2,4-D was applied to reservoirs for aquatic weed

control, about half of water samples from within treatment

areas contained 2,4-D, and the highest concentration

(0.027 mg per liter) was less than half of the HAL (0.070

mg per liter). Newton and others (1994) aerially applied

glyphosate at three times the normal forestry usage rate [4

pounds ai per acre (4.4 kg per hectare)], no buffers were left,

and all streams and ponds were sprayed. Initial water

concentrations were 0.031 and 0.035 mg per liter in Oregon

and Georgia, respectively, and 1.237 mg per liter in Michigan on the day of application. After day one, glyphosate

concentrations dropped to below 0.008 mg per liter on all

three sites for the duration of the study. The HAL was

exceeded on only one of three sites and then for only 1 day.

There is little information on the movement of metsulfuron

to streams. Michael and others (1991) found trace residues

of metsulfuron in shallow monitoring wells in Florida where

24 wells were sampled to a depth of 6 feet (1.8 meters).

Metsulfuron was detected (0.002 mg per liter) in 1 of 207

samples collected during 2 months after application.

Pesticides movement into streams is well documented, but

movement into ground water is not well researched. During

movement from streamwater into ground water, concentrations should be reduced considerably for several reasons.

Infiltrating pesticides must pass through several physical

barriers or layers before reaching ground water. As they pass

Table 13.3—Frequency and occurrence of surface and ground water contamination from pesticide use in North Americaa

(continued)

Water

Pesticide type Location Maximum Range Comments Reference

- - - - - Milligrams per liter - - - - -

Picloram

 (cont.) G Saskatchewan,

Canada 0.000225 NR Natural spring flow Wood and Anthony 1997

G Iowa, Maine, .049 0.00063 – 0.049 Fewer than 2% of well samples Funari and others 1995

Minnesota, were positive

North Dakota

Strychnine NR NR NR NR NR

Thiram NR NR NR NR NR

Triclopyr S Florida .002 NG Coastal Plain flatwoods catch- Bush and others 1988

ments near Gainesville, FL

S Ontario, Canada .35 .23 – .35 Intentional aerial application Thompson and others 1991

to boreal forest stream

Zinc

phosphide NR NR NR NR NR

Ai = active ingredient; G = ground water; ND = not detected; NG = not given; NR = no reports found in published literature; S = surface water;

smz = streamside management zone; USFS = USDA Forest Service.

aThis table summarizes the levels of pesticides reported in the literature at specific sites and is representative of the literature from North America. However, it

cannot be extrapolated for purposes of prediction.

b The authors do not provide specific location.

Pesticides
147

through each layer, they are degraded, diluted, and metabolized. Surface water provides a medium for dilution,

hydrolysis, and photolysis. Aquatic vegetation can metabolize pesticides. Microbes associated with coarse and fine

particulate organic matter found naturally in streams also

metabolize pesticides.

In order for water on the soil surface to carry pesticides into

ground water, it must pass through the soil column. Here

again, processes work to reduce the potential for pesticides

to reach ground water. Pesticides percolating through the

soil column are adsorbed to soil particles, reducing the

amount reaching the ground water. Pesticides adsorbed onto

soil particles may be irreversibly bound, released slowly, or

further metabolized by microbes. Once pesticides reach

ground water, they may degrade further. Cavalier and others

(1991) found that microbes degraded herbicides, including

2,4-D, in ground water.

Thus, ground water concentrations of pesticides should be

considerably lower than observed in surface water. Funari

and others (1995) reviewed the literature and reported the

range of maximum ground water concentrations of pesticides, including those used in forestry, agriculture, home

and garden, and on industrial rights-of-way. The maximum

range of values for 2,4-D (0.0002 to 0.0495 mg per liter),

hexazinone (0.009 mg per liter), and picloram (0.00063 to

0.049 mg per liter) are much lower than the HAL’s for those

compounds.

The National Water-Quality Assessment Program

(NAWQA) conducted by the U.S. Geological Survey began

in 1991. The focus of NAWQA is to identify nutrient and

pesticide contamination of water throughout the United

States. The 1999 NAWQA report (found at http://water.usgs.

gov/pubs/circ/circ1225/index.html) makes little mention of

forest sites or forestry pesticides, but concludes that:

“Concentrations of nutrients and pesticides in streams and

shallow groundwater generally increase with increasing

amounts of agricultural and urban land in a watershed.” The

report focused on more than 50 major river basins and

aquifers supplying water to more than 60 percent of the

population and approximately half of the area of the United

States. Few forestry pesticides other than 2,4-D were found

in these basins or aquifers.

Even in predominantly agriculture areas, the report states:

One of the most striking results for shallow groundwater in agricultural areas, compared with streams, is

the low rate of detection for several high-use herbicides other than atrazine. This is probably because

these herbicides break down faster in the natural

environment compared to atrazine.

Atrazine is principally used in growing corn (maze). It has

not been used on NFS land since 1992. While not directly

addressing forestry pesticides and drinking water, these

NAWQA conclusions support the above research findings

and conclusions that ground water contamination by

pesticides should be lower than observed for surface water.

Because surface water contamination from forest sites

treated according to label directions does not exceed HAL’s,

it is very unlikely that ground water contamination would

exceed HAL’s.

Several of the pesticides in table 13.2 have not been

reported in water. They include chloropicrin, chlopyralid,

dazomet, and thiram. Chloropicrin and dazomet are soil

fumigants, which are gases in their active form and are used

only for seedling production. Chlopyralid is a relatively new

compound in the United States. Thiram is a dimethyl

dithiocarbamate fungicide, principally used in forestry for

seed protection.

There is very little water-quality data for pesticides used in

nursery disease control and soil fumigation. More than 71

percent of fungicides and fumigants used on NFS land are

applied in nurseries. Intense use in a nursery may result in

localized ground water contamination. Three pesticides

(chloropicrin, dazomet, and methyl bromide) make up this

group of intensively used agents. Chloropicrin is toxic to

plants and is used in combination with other chemicals for

fumigating seedbeds. Dazomet, a soil fumigant, is a gas and

is relatively insoluble in water (3 grams per liter). However,

dazomet is unstable in water and quickly breaks down into

methyl isothiocyanate (MITC), formaldehyde, monomethylamine, and hydrogen sulfide. All are toxic, but the

most toxic is MITC. The RfD for formaldehyde is 0.2 mg

per kilogram per day. However, EPA has classified formaldehyde as a compound of medium carcinogenic hazard to

humans. Methyl bromide is very toxic. Data are insufficient

to determine whether frequent use of these three pesticides

adversely impacts water quality, either locally or over an

expanded area.

Reliability and Limitation of Findings

Most data reviewed in this chapter come from scientific

literature. The data listed in table 13.3 and derived from

Larson and others (1997) were extracted from in-house

reports from the U.S. Geological Survey, the EPA, State,

and local governmental departments for the environment

and scientific literature. Reports published in scientific

literature are the most reliable because they were subject

to peer review and scrutiny for validity of methods,

Chapter 13
148

completeness of data, and interpretation of the data.

Monitoring data from in-house publications and reports may

be less reliable.

Some variability in results of individual studies is due to

regional soil and climate differences. In the South, infiltration rates on many forestry sites are generally low, owing to

the highly eroded condition of the soils. Here, precipitation

intensity frequently exceeds infiltration rates, producing

overland flow on newly prepared sites. Overland flow may

lead to much higher pesticide concentrations in stormflow

than in other areas of the country with much higher infiltration rates. Very high infiltration rates are typical of soils in

the Pacific Northwest. Therefore, if streams are protected by

buffers, broadcast application of pesticides generally results

in stream contamination either via direct application or

through baseflow contributions. In general, levels of

contamination are lowest where infiltration rates are highest.

Care must always be exercised in extrapolating data from

local studies on drinking water to a regional or larger scale.

However, three strategies of worst-case scenarios used in

these studies mitigate against high levels of uncertainty:

(1) several studies have investigated the impacts of pesticides applied directly to surface water; (2) several studies

have investigated the impacts on water of pesticides applied

at several times the prescribed rate; and (3) most of the

studies conducted specifically on forestry sites treated the

entire catchment from which water samples were taken,

resulting in samples with levels of pesticide contamination

greater than are likely to occur anywhere downstream.

Research on the impacts of pesticides applied directly to

surface water used the worst-case scenario for forest

operational treatments in which pesticide was applied at

normal rates directly to surface water (ponds and streams).

These studies did not find any contamination of water at

levels above the HAL for any pesticide studied. Research on

aquatic impacts from pesticides applied at several times the

labeled rate used the worst-case scenario for operational

treatments where an area might receive multiple

applications in error or where small spills occurred. In these

studies, HAL’s were exceeded by only a few percent and

only briefly, usually for less than a few hours. Both worstcase scenarios just described were combined with the third

worst-case scenario in which all sampling was conducted on

surface water found within the treated area. In this case,

most of the water was in small pools or ephemeral to firstorder streams. While ephemeral to first-order streams or

pools are unlikely to be drinking water sources because of

low yield, they do represent water most likely to be severely

contaminated from normal forest pesticide applications.

Even these waters were not contaminated at levels

exceeding HAL’s, except where pesticide was applied at

several times the labeled rate as described above.

In addition, data on contamination of water for the pesticides in table 13.3 have been taken from a number of studies

conducted in North America and the findings are generally

similar. These studies have, with a few exceptions, confirmed the absence of significant contamination of drinking

water. The exceptions were those cases in which a pesticide

was applied directly to water, and the high concentrations

observed in those studies were at or only slightly above

drinking water standards. These high concentrations lasted

only a few hours at most before dropping well below current

HAL’s. It is clear from the available literature that use of

pesticides in strict accordance with label directions on NFS

land cannot be expected to contribute significantly to ground

water or drinking water contamination. It is also clear that

pesticides, unless clearly labeled for aquatic uses, must not

be applied directly to water, and that pesticides should be

used around water resources, which are particularly sensitive only after careful consideration of the ramifications.

Limitations of the data are obvious for the few chemicals

that have not been investigated. We need data on them as

well as other chemicals about which little information is

available. Additional limitations include lack of sufficient

testing for health effects as indicated in table 13.2. The

question of cumulative toxicological effects has not been

addressed for any of the pesticide mixtures utilized in

modern forest management.

Research Needs

Several issues related to vegetation management need

additional research.

1. One issue is impacts of frequent, repeated use of fungicides and fumigants in nursery operations on nearby

water quality.

2. Another issue is effectiveness of buffer width and

composition. There is too little information on the

processes and interactions of site-specific characteristics

with pesticide chemistry that permit buffers to mediate

against contamination of streams and surface waters in

general. These processes and the interactions of pesticide

chemistry with site-specific conditions must be identified

and understood so managers can design and install

optimally functional buffers to protect the water resource

and its associated aquatic ecosystem.

Pesticides
149

3. A third issue is that several pesticides in table 13.3 lack

published reports relating their use to occurrence in

water.

4. Still another issue is the effects of commonly used

pesticide mixtures, as opposed to single compounds, on

the water resource.

Key Points

Relative to agricultural, urban, and other uses of pesticides,

very small amounts are used on NFS land. Further, the use

patterns for any specific piece of land are infrequent, except

in the case of vegetation protection from pathogens,

animals, and insects where annual treatment may be

required, especially in greenhouses and nurseries. Pesticides

used on NFS land are also used around the World to

accomplish management goals similar to those on NFS land,

but often in a much more intensive way. Even with the

widespread use of pesticides in North America, those

typically used on NFS land have not been identified in

surface or ground water at sufficiently high concentrations

as to cause drinking water problems. Their rapid break down

by physical, chemical, and biological routes coupled with

use patterns precludes the development of water contamination problems unless they are applied directly to water. Even

though these same pesticides are used around homes, in

urban and in agricultural settings, their use in forest management is still controversial in the public arena. Therefore,

their use should be carefully planned and all agency, local,

State, and Federal laws should be followed. It is especially

important to follow all label directions because pesticide

labels are legal documents specifying Federal laws pertaining to their use. Best management practices should be

carefully adhered to and use around drinking water supplies

should be avoided, except where permitted by the label.

Wherever pesticides are used, precautions should always be

taken to protect drinking water sources from contamination.

Literature Cited

Bouchard, D.C.; Lavy, T.L.; Lawson, E.R. 1985. Mobility and persistence of hexazinone in a forest watershed. Journal of Environmental

Quality. 14: 229–233.

Bruce, B.W.; McMahon, P.B. 1996. Shallow ground-water quality beneath

a major urban center. Journal of Hydrology. 186: 129–151.

Bush, P.B.; Michael, J.L.; Neary, D.G.; Miller, K.V. 1990. Effect of

hexazinone on groundwater quality in the Coastal Plain. Proceedings of

the Southern Weed Science Society; 1990 January 15–17; Atlanta.

Champaign, IL: Southern Weed Science Society. 43: 184–194.

Bush, P.B.; Neary, D.G.; Taylor, J.W. 1988. Effect of triclopyr amine and

ester formulations on groundwater and surface runoff water quality in

the Coastal Plain. Proceedings of the Southern Weed Science Society;

1988 January 18–20; Tulsa, OK. Champaign, IL: Southern Weed Science

Society. 41: 226–232.

Cavalier, T.C.; Lavy, T.L.; Mattice, J.D. 1989. Assessing Arkansas

ground water for pesticides: methodology and findings. Ground Water

Monitoring Review. 9: 159–166.

Cavalier, T.C.; Lavy, T.L.; Mattice, J.D. 1991. Persistence of selected

pesticides in ground-water samples. Ground Water. 29: 225–231.

Comerford, N.B.; Neary, D.G.; Mansell, R.S. 1992. The effectiveness of

buffer strips for ameliorating offsite transport of sediment, nutrients, and

pesticides from silvicultural operations. Tech. Bull. 631. Research

Triangle Park, NC: National Council of the Paper Industry for Air and

Stream Improvement. 48 p.

Edwards, W.M.; Triplett, G.B., Jr.; Kramer, R.M. 1980. A watershed

study of glyphosate transport in runoff. Journal of Environmental

Quality. 9(4): 661–665.

Feng, J.C.; Thompson, D.G.; Reynolds, P.E. 1990. Fate of glyphosate in

a Canadian forest watershed. 1: Aquatic residues and off-target deposit

assessment. Journal of Agricultural and Food Chemistry. 38(4):

1110–1118.

Funari, E.; Donati, L.; Sandroni, D.; Vighi, M. 1995. Pesticide levels in

groundwater: value and limitations of monitoring. In: Vighi, M.; Funari,

E., eds. Pesticide risk in groundwater. Boca Raton, FL: Lewis Publishers. 275 p.

Holmes, S.B.; Millikin, R.L.; Kingsbury, P.D. 1981. Environmental effects

of a split application of Sevin–2–oil. Rep. FPM–X–46. Sault Ste. Marie,

Ontario: Canadian Forestry Service. 58 p.

Johnsen, T.N., Jr. 1980. Picloram in water and soil from a semiarid

pinyon-juniper watershed. Journal of Environmental Quality. 9:

601–605.

Kolpin, D.W.; Kalkhoff, S.J.; Goolsby, D.A. [and others]. 1997.

Occurrence of selected herbicides and herbicide degradation products in

Iowa’s ground water, 1995. Ground Water. 35: 679–688.

Koterba, M.T.; Banks, W.S.L.; Shedlock, R.J. 1993. Pesticides in

shallow groundwater in the Delmarva Peninsula. Journal of Environmental Quality. 22: 500–518.

Larson, S.J.; Capel, P.D.; Majewski, M.S. 1997. Pesticides in surface

water: distribution, trends, and governing factors. Ser. 5.3: Pesticides in

the hydrologic system. Chelsea, MI: Ann Arbor Press. 373 p.

Leveille, P.; Legris, J.; Couture, G. 1993. Results of spot checks in lotic

environments after spraying of glyphosate in forests from 1989 to 1991.

Quebec, Canada: Ministere des Forets du Quebec. 21 p.

Michael, J.L.; Neary, D.G. 1993. Herbicide dissipation studies in southern

forest ecosystems. Environmental Toxicology and Chemistry. 12:

405–410.

Michael, J.L.; Neary, D.G.; Fischer, J.B.; Gibbs, H.G. 1991.

Metsulfuron in surface groundwater of a north Florida flatwoods.

Proceedings of the Southern Weed Science Society; 1991 January

14–16; San Antonio, TX. Champaign, IL: Southern Weed Science

Society. 44: 233–237.

Chapter 13
150

Michael, J.L.; Neary, D.G.; Wells, M. 1989. Picloram movement in soil

solution and streamflow from a Coastal Plain forest. Journal of

Environmental Quality. 18: 89–95.

Michael, J.L.; Webber, E.C.; Bayne, D.R. [and others]. 1999.

Hexazinone dissipation in forest ecosystems and impacts on aquatic

communities. Canadian Journal of Forest Research. 29: 1170–1181.

Miller, J.H.; Bace, A.C. 1980. Streamwater contamination after aerial

contamination of a pelletized herbicide. Res. Note SO–255. New

Orleans: U.S. Department of Agriculture, Forest Service, Southern

Forest Experiment Station. 4 p.

Neary, D.G.; Bush, P.B.; Douglas, J.E.; Todd, R.L. 1985. Picloram

movement in an Appalachian hardwood forest watershed. Journal of

Environmental Quality. 14: 585–592.

Neary, D.G.; Bush, P.B.; Grant, M.A. 1986. Water quality of ephemeral

streams after site preparation with the herbicide hexazinone. Forest

Ecology and Management. 14: 23–40.

Newton, M.; Horner, L.M.; Cowell, J.E. [and others]. 1994. Dissipation

of glyphosate and aminomethylphosphonic acid in North American

forests. Journal of Agricultural and Food Chemistry. 42: 1795–1802.

Norris, L.A. 1975. Dicamba residues in streams after forest spraying.

Bulletin of Environmental Contamination and Toxicology. 13: 1–8.

Norris, L.A. 1981. Phenoxy herbicides and TCDD in forests. Residue

Reviews. 80: 65–135.

Pimentel, D.; Levitan, L. 1986. Pesticides: amounts applied and amounts

reaching pests. BioScience. 36(2): 86–91.

Smith, N.J.; Martin, R.C.; St. Croix, R.G. 1996. Levels of the herbicide

glyphosate in well water. Bulletin of Environmental Contamination and

Toxicology. 57: 759–765.

Stanley, R.A.; Bates, A.L.; Bedsole, J.A.; Hall, T.F. 1974. Program

diversity and control of Eurasian watermilfoil in Tennessee Valley

reservoirs. In: Proceedings of the Weed Science Society of America;

1974 February 12–16; Las Vegas, NV: Champaign, IL: Southern Weed

Science Society. 19 p.

Sundaram, K.M.S.; Szeto, S.Y. 1987. Distribution and persistance of

carbaryl in some terrestrial and aquatic components of a forest

environment. Journal of Environmental Science and Health, B

(Pesticides, Food Contaminants and Agricultural Wastes). B22:

5, 579–599.

Thompson, D.G.; Staznik, B.; Fontaine, D.D. [and others]. 1991. Fate

of triclopyr ester released in a boreal forest stream. Environmental

Toxicology and Chemistry. 10(5): 619–632.

U.S. Department of Agriculture, Forest Service. 1998. Report of the Forest

Service fiscal year 1997. Washington, DC: U.S. Department of Agriculture, Forest Service. 147 p.

U.S. EPA. 1988a. Picloram health advisory. Washington, DC: U.S.

Environmental Protection Agency, Office of Drinking Water. 18 p.

U.S. EPA. 1988b. U.S. EPA integrated risk information system substance

file–Ally. CASRN 74223–64–6. Washington, DC: U.S. Environmental

Protection Agency. [Not paged]. [Available on the Web at: http://

www.epa.gov/ngispgm3/iris/subst/0288.htm].

U.S. EPA. 1989. Drinking water health advisory. Pesticides. U.S.

Environmental Protection Agency, Office of Drinking Water Health

Advisories. Chelsea, MI: Lewis Publishers, Inc. 819 p.

U.S. EPA. 1990. Drinking water regulations and health advisories.

Washington, DC: U.S. Environmental Protection Agency, Office of

Drinking Water. 10 p.

U.S. EPA. 1992. U.S. EPA integrated risk information system substance

file — Thiram. CASRN 137–26–8. Washington, DC: U.S. Environmental Protection Agency. [Not paged]. [Available on the Web at: http://

www.epa.gov/ngispgm3/iris/subst/0267.htm].

U.S. EPA. 1993. Health advisories for drinking water contaminants. U.S.

Environmental Protection Agency, Office of Water Health Advisories.

Boca Raton, FL: Lewis Publishers. 261 p.

U.S. EPA. 1996. Hexazinone health advisory. Washington, DC: U.S.

Environmental Protection Agency, Office of Drinking Water. 24 p.

U.S. EPA. 1997. Imazapyr (Arsenal) pesticide tolerances: final rule.

Pesticide management education program. Ithaca, NY: Cornell

University. [Not paged]. [Available on the Web at: http://

pmep.cce.cornell.edu/profiles/herb-growthreg/fatty-alcohol-monuron/

imazapyr/tol-imazapyr.html#top].

U.S. EPA. 1998a. Draft Resource Conservation and Recovery Act; waste

minimization PBT chemical list; app. C: Waste minimization

prioritization tool human noncancer scores. Washington, DC: U.S.

Environmental Protection Agency, Office of Solid Waste, Waste

Minimization Branch. [Not paged]. [Available on the Web at: http://

www.epa.gov/epaoswer/hazwaste/minimize/chemlist/ap_c.pdf].

U.S. EPA. 1998b. Reregistration eligibility decision (R.E.D.) facts:

Triclopyr. EPA–738–F–98–007. Washington, DC: U.S. Environmental

Protection Agency, Office of Prevention, Pesticides and Toxic Substances (7508C). [Not paged]. [Available on the Web at: http://

www.epa.gov:80/oppsrrd1/REDs/factsheets/2710fact.pdf].

Wood, J.A.; Anthony, D.H.J. 1997. Herbicide contamination of prairie

springs at ultratrace levels of detection. Journal of Environmental

Quality. 26: 1308–1318.

Pesticides
151

Part IV:

Effects of Grazing

Animals, Birds, and

Fish on Water Quality

Little green herons (Butorides virescens) are common to most water areas.

Photo by Bill Lea
153

Chapter 14

Domestic Grazing

John C. Buckhouse1

Introduction

Livestock grazing is a significant use of rangeland, and

grazing practices can affect the quality of public drinking

water sources. In general, activities that provide healthy

range usually help maintain—or at least do not significantly

degrade—water quantity and quality for domestic use.

Rangeland management is not simply another phrase for

livestock management. Rangeland often has climate, soils,

topography, and precipitation characteristics or all that are

unsuitable for intensive agriculture without irrigation or

other intensive managerial inputs. Many rangelands have

trees and are grazed.

Western rangelands are emphasized because most livestock

grazing on natural forests and grasslands occurs in the West.

Some also occurs in other regions. The publications cited

here are meant to reflect physical and biological principles

associated with water quality and livestock use. Therefore,

while specific effects of livestock or water quality may vary,

the underlying principles should be consistent.

Erosion and Sedimentation

Erosion and its consequence, sedimentation, are generally

considered to be the number one problem associated with

wildland watershed management.

Issues and Risks

Overgrazing weakens or kills vegetation, reducing soil cover

and thereby accelerating surface erosion. With increased

erosion, soil fertility declines and sediment yields increase.

When significant amounts of sediment enter stream channels from rangeland, channels destabilize and widen,

creating additional sources of sediment (U.S. Department of

the Interior, Bureau of Land Management 1993, 1998).

Sediment reduces water clarity and the oxygen-carrying

capacity of the stream. Nutrients attached to sediment

heighten the possibility of eutrophication. See chapter 2 for

more discussion of sediment impacts on drinking water

quality.

Findings from Studies

Considerable research is available on the relationships

between livestock grazing and erosion and sedimentation

(fig. 14.1). Several textbooks summarize the effects of

livestock numbers, livestock types, timing of grazing, and

animal distribution on vegetation and erosion (Holecheck

and others 1989, Stoddart and others 1975). At high densities, grassland vegetation promotes production of soil

organic material and increases infiltration rates (Buckhouse

and Gaither 1982, Buckhouse and Mattison 1980). Therefore, grazing should be managed to maintain the density of

vegetation, reducing erosion, and sediment yields.

In the riparian zone, the relationship of livestock grazing to

streambank erosion has been studied (Bohn and Buckhouse

1985, Buckhouse and Bunch 1985, Buckhouse and others

1981). These researchers found it is possible to manage

livestock grazing in ways that enhance riparian vegetation

and protect streambanks. Grazing throughout the growing

season harms vegetation and results in increased streambank

sloughing. Conversely, grazing that is timed to accommodate plant growth and physiology can have positive effects

on streams and the quality of water in them. In Oregon, for

example, early-season grazing enhanced riparian shrub

growth (Buckhouse and Elmore 1997).

J.M. Skovlin (1984) prepared an excellent review of

livestock grazing research. He described effects on vegetation, streambank stability, and various aspects of water

quality. He also compiled a large reference list and provided

recommendations and prescriptions for grazing management. This document is an exhaustive look at many aspects

of grazing management.

Reliability and Limitations of Findings

This work has been repeated and verified in several locations by several researchers. It is known that loss of plant

1

 Professor, Oregon State University, Department of Rangeland Resources,

Corvallis, OR.
154

Figure 14.1—Potential sediment production in 10 Blue Mountain

ecosystems in Oregon. Ecosystems: L = larch, M = meadow,

LP = lodgepole pine, DF = douglas-fir, A = alpine, PP = ponderosa pine,

SF = spruce-fir, G = grassland, SA = sagebrush, J = juniper (Buckhouse

and Gaither 1982). Different lower-case letters indicate differences in

statistical significance (P<0.10).

cover increases sediment yield. The key to applying findings

is to think in terms of plant physiology and plant response to

grazing. The goal of proper grazing is to maintain vegetation

and soil organic material, preserving the soil’s ability to

infiltrate water, resist erosion, and store and slowly release

precipitation.

Managerial success depends on understanding ecosystems,

the vegetation present or desired, and management objectives, and then prescribing herbivory appropriate to the site.

The Forest Service and the U.S. Department of the Interior,

Bureau of Land Management, have cooperated to develop

Proper Functioning Condition (PFC), a first-cut methodology to monitor wildland streams. Wayne Elmore and his

riparian team have based their approaches on these concepts, which have been widely adopted (U.S. Department of

the Interior, Bureau of Land Management 1998). However,

there have been no systematic studies of how effective these

methods are for protecting the quality of public drinking

water sources.

Research Need

Refinement of the relationships among herbivory, vegetation, and sediment production on specific sites in specific

ecosystems needs more research.

Bacteria and Protozoa

Waterborne, pathogenic bacteria and protozoa have long

been recognized as causes of human disease. Many diseases

can be transmitted only among members of the same

species, but a significant number can be transmitted to hosts

of different species. The concern is that contamination of the

drinking water supplies by humans and by a variety of

animal species poses hazards to human health.

Issues and Risks–Fecal Bacteria

Escherichia coli is a ubiquitous bacterium found in the gut

of all warm-blooded animals. It is used as an indicator

species in bacteriological testing. Other pathogenic

organisms are difficult to trap, difficult to analyze, and

expensive to process. As a consequence, testing for fecal

coliform bacteria has become the accepted surrogate

sampling protocol, with the understanding that if fecal

coliforms are present, then pathogens are potentially

present. Escherichia coli was long considered to be benign.

However, in recent years several pathogenic strains of E.

coli have developed and gained national attention due to the

health risk they pose.

Findings from Studies (Bacteria)

Walter and Bottman (1967) reported on the two tributaries

of the watershed supplying the city of Bozeman, MT, with

its drinking water. One tributary was fenced and human

activity was limited. A corresponding tributary was not

restricted from human entry. The fenced watershed consistently had higher fecal bacteria loads than did the open

watershed. After considerable monitoring and study, it was

realized that the closed watershed had become a de facto

refuge for wildlife. The increased animal use resulted in the

higher fecal bacteria numbers in streamwater.

Coltharp and Darling (1973) studied three neighboring

watersheds with different combinations of animals grazing

and browsing: wildlife, sheep and wildlife, and cattle and

wildlife. They found the lowest numbers of bacteria in the

streams in the wildlife-only watershed. The sheep-andwildlife watershed was next lowest, and the cattle-andwildlife watershed had the highest bacteria counts. There

was high statistical variation in the numbers, but in all three

strategies, the bacterial counts were <100 per 100 milliliters

(ml). The authors attributed the differences to animal

numbers and distribution. The cattle tended to congregate

near the stream, while the sheep were herded and spent

more time in the uplands.

Bohn and Buckhouse (1981, 1983) sampled water in

northeastern Oregon and found similar results. On watersheds with a resident deer and elk population, where grazing

by livestock was excluded for several months during the

winter, the bacteria numbers were low (10 to 20 per 100 ml).

At the end of the summer, after livestock had grazed the

area for several months, the bacteria numbers were slightly

elevated (20 to 40 per 100 ml).

Domestic Grazing

 0

 500

1,000

1,500

2,000

L M LP DF A PP SF G SA J

Potential sediment production

(kg/ha)

a

15 a

22

ab

36

ab

109

ab

140

ab

183

ab

217

b

431

c

1284

c

1572
155

Research in central Oregon studied the fate of the coliform

bacteria (Biskie and others 1988; Larsen and others 1988,

1994; Moore and others 1988; Sherer and others 1992). In

this Great Basin rangeland, they discovered that statistically

only 5 days per year experienced enough precipitation to

produce overland flows. Consequently, the probability of

washing fecal material into streams was relatively low.

Cows defecated approximately 11 times a day, but less than

one of these defecations landed in the stream directly or

within 1 meter of the stream where it might be washed into

the stream. Further experiments were conducted to determine the fate of bacteria defecated directly into the stream.

Ninety percent of these bacteria precipitated to the stream

bottom and attached to sediments. Sediment samples

collected over the next several weeks showed that 90

percent of the lodged bacteria had died within 40 days

(Biske and others 1988; Larsen and others 1988, 1994;

Moore and others 1988; Sherer and others 1992).

The same team of researchers studied the effect of a

strategically placed watering trough on livestock use of a

flowing stream in both summer and winter (Clawson and

others 1994, Miner and others 1992). When snow covered

the ground, cattle were fed hay, and 95 percent of them used

the ground water-fed trough, as opposed to 5 percent that

used the stream. It is speculated that the warmer ground

water (approximately 50 °F or 10 °C) held much greater

appeal to the cattle than did the 32 °F (0 °C) creek. In the

summer, riparian zones provided lush vegetation, shade, and

water. The trough relieved some impact during the summer,

but not as dramatically as during the winter. In late summer,

when most of the available vegetation near the riparian zone

had been consumed, approximately 25 percent of the

livestock drank from the trough, while the remaining 75

percent preferred the stream.

Findings from Studies (Protozoa)

Giardia and Cryptosporidium have drawn considerable

attention recently. Both are debilitating to humans and can

be carried by a wide variety of warm-blooded animals from

waterfowl to rodents; from deer, elk, and beaver to livestock

(see appendix B and chapter 15). Both Giardia and

Cryptosporidium have been known for decades, but only

recently have routine testing following gastrointestinal

complaints from citizens been conducted. It is probable that

what is now diagnosed as Cryptosporidium would at an

early time have been seen simply as flu or food poisoning.

Cryptosporidium oocysts have been found in association

with both wild and domestic animals. Calves consistently

shed greater numbers of oocysts than do older animals

(Atwill 1996). Apparently by 4 months of age, calves

develop a resistance and the number of oocysts shed is

dramatically reduced. An alternative to livestock exclusion

from areas where Cryptosporidium may be a concern is to

ensure that livestock grazing the watershed are older than

4 months. The relationship between oocysts and age of wild

animals is unknown and represents a research question.

Reliability and Limitations of Findings

The fecal bacteria research is reliable and has been consistent over time. The Cryptosporidium work, while compelling, is relatively recent.2 Tate and his colleagues at the

University of California, Davis, continue to investigate these

relationships, the biology, and the management of

Cryptosporidium. Conceptually, these experiments were

solid. Site-specific relationships dictate specific biological

responses. However, regional climatic and temperature

differences may influence the range of variability inherent in

these relationships and to date little is known about wild

animals except that they carry the organisms and shed them

in their feces. Water from wildlands including wilderness,

can contain pathogens that cause human disease if the water

is not adequately treated for purification.

Research Needs

Cryptosporidium research is still in its infancy. Further

studies dealing with the origin and fate of the organisms are

needed. Very little is known about pathogenic E. coli in

livestock under range conditions, and further investigation is

needed.

Chemical and Nutrient Impacts

A number of chemical compounds are associated with

sediment transport. Rangelands are sometimes treated with

herbicide or fertilizer compounds, but the primary chemical

or nutrient problems are with phosphate (PO4

+3) and nitrate

(NO3

-1) associated with eroding soils and fecal material.

Issues and Risks

Cycling of phosphate and nitrate are essential to plant

growth. When vegetation is grazed, potential problems exist

where phosphates and nitrates from feces or attached to

Chapter 14

2

 Personal communication. 1999. Kenneth Tate, Professor, Department

Agronomy and Range Science, University of California, Davis, CA 95616.
156

erosion particles that find their way to a stream. Excessive

levels of these nutrients in streams can produce algal blooms

and eutrophication. See chapter 2 for discussion of nutrient

impacts on drinking water.

Findings from Studies

Frequently phosphate and nitrate reach streams in association with sediments. See chapter 3, which discusses overland flow, erosion, and sediment transfer.

When fecal material enters a stream, phosphate and nitrate

concentrations rise, but wetlands can reduce these concentrations. On the Wood River system in Oregon, a nutrient

loading was studied on streams originating from the Crater

Lake National Park, traversing national forest, then private

grazing land, and emptying into a lake listed as hypoeutrophic. The concern had been that nutrient loading would

be exacerbated when the water flowed through the grazed

land due to fecal contamination. The data did not bear out

this fear. In fact, the phosphate and nitrate levels decreased.3

It was speculated that the wetlands in the system represented

a natural nutrient sink. In wetlands, chemical processes

associated with anaerobic conditions reduced phosphate and

nitrate concentrations. Furthermore, wetland plants take up

nutrients from the aqueous system. If animals eat wetland

vegetation, nutrients are consumed. Thus, these nutrients are

not redistributed back into the water as when the plant

senesces in place and dies. Apparently the livestock grazing

the wetland consumed the vegetation and its nutrients, and

later redistributed the nutrients away from the stream in

their feces. The result was the observed decline in waterborne nutrient concentrations.

Reliability and Limitations of Findings

The relationship between nutrient loading and sediment is

clear. The relationship between livestock grazing of wetlands and the possibility of nutrient loading abatement is not

proven. The relationships among erosion, sedimentation,

and nutrient loading are universal and are bases for erosion

control worldwide. The possibility of using livestock for

biological control of weeds, for improving plant communities, for promoting species that encourage infiltration and

reduce overland flow, and even for reducing nutrient loading

(Bedell and Borman 1997) are based on solid research and

management experience. These concepts are predicated on

the ability of vegetation and organic material to enhance

infiltration, which reduces overland flows and subsequent

erosion. The concept of using livestock to harvest nutrientrich wetland vegetation is logical but not tested.

Research Needs

Further research is needed on the relationships between

livestock harvesting of streamside vegetation and nutrient

loading of streams.

Key Points

In rangeland, grazing can affect drinking water. Water

quality can be protected by nurturing upland and riparian

vegetation, which can increase the soil’s infiltration capacity

and reduce surface runoff. By understanding the relationships between plant physiology and animal herbivory, one

can tailor grazing practices to enhance infiltration. Both

timing and intensity of grazing must be managed. Increased

infiltration reduces sediment yield, bacterial counts, nutrient

concentrations, and water temperatures. Published data

clearly indicate that improper, abusive grazing can degrade

the quality of public drinking water sources. It is also clear

that proper, prescriptive grazing can produce positive

environmental benefits, and research is needed to develop

and test methods to accomplish them.

Literature Cited

Atwill, E.R. 1996. Assessing the link between rangeland cattle and

waterborne Cryptosporidium parvum infection in humans. Rangelands.

18: 48–51.

Bedell, T.E.; Borman, M.M. 1997. Watershed management guide for the

interior Northwest. Ext. Man. 8436. Corvallis, OR: Oregon State

University Extension Service. 84 p.

Biskie, H.A.; Sherer, B.M.; Moore, J.A. [and others]. 1988. Fate of

organisms from manure point loading into a rangeland stream. Pap.

88–2081. American Society of Agricultural Engineers: 1–14.

Bohn, C.C.; Buckhouse, J.C. 1981. Interpreting coliform counts in

northeast Oregon rangeland streams. Spec. Prog. Rep. 620. Research in

rangeland management. Corvallis, OR: Oregon State University,

Agricultural Experiment Station: 15–17.

Bohn, C.C.; Buckhouse, J.C. 1985. Coliforms as an indicator of water

quality in wildland streams. Journal of Soil and Water Conservation.

40: 95–97.

Buckhouse, J.C.; Bohn, C.C. 1983. Response of coliform bacteria

concentration to grazing management. Spec. Prog. Rep. 682. Corvallis,

OR: Oregon State University, Agricultural Experiment Station: 1–8.

Domestic Grazing

3

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of exclosures on Mill and McKay Creeks: a validation of management.

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within various vegetative communities found in the Blue Mountains of

Oregon. Journal of Soil and Water Conservation. 37: 120–122.

Buckhouse, J.C.; Knight, R.W.; Skovlin, J.M. 1981. Streambank erosion

and ungulate grazing relationships. Journal of Range Management. 34:

339–340.

Buckhouse, J.C.; Mattison, J.L. 1980. Potential soil erosion of selected

habitat types of the high desert region of central Oregon. Journal of

Range Management. 33: 282–286.

Clawson, J.; Buckhouse, J.C.; Miner, J.R.; Moore, J.A. 1994. An

evaluation of off-site water development and fence configurations

designed to ameliorate water quality impacts by livestock [Abstract].

Corvallis, OR: Oregon State University, Oregon Academy of Science.

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symposium on rural environmental engineering; 1973 September 15;

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principles and practices. Englewood Cliffs, NJ: Prentice-Hall. 501 p.

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quality benefits of having cattle manure deposited away from streams.

Bioresource Technology. 48: 113–118.

Miner, J.R.; Buckhouse, J.C.; Moore, J.A. 1992. Will a water trough

reduce the amount of time hay-fed livestock spend in the stream (and

therefore improve water quality)? Rangelands. 14: 35–38.

Moore, J.A.; Buckhouse, J.C.; Miner, J.R. 1988. Water quality impacts

of alternate rangeland management systems. Comp. Rep. 85–CSRS–2–

2718. [Location of publisher unknown]: U.S. Department of Agriculture,

Cooperative State Research Service. 125 p.

Odum, E.P. 1971. Fundamentals of ecology. Philadelphia: W.B. Saunders

Co. 574 p.

Sherer, B.M.; Miner, J.R.; Moore, J.A.; Buckhouse, J.C. 1988.

Resuspending organisms from a rangeland stream bottom. Transactions

of the American Society of Agricultural Engineers. 31: 1217–1222.

Sherer, B.M.; Miner, J.R.; Moore, J.A.; Buckhouse, J.C. 1992. Indicator

bacterial survival in stream sediments. Journal of Environmental Quality.

21: 591–595.

Skovlin, J.M. 1984. Impacts on wetlands and riparian habitat: a review of

our knowledge. In: Developing strategies for rangeland management. A

report prepared by the Committee on Developing Strategies for

Rangeland Management, National Research Council, National Academy

of Sciences. Boulder, CO: Westview Press: 1001–1104.

Stoddart, L.A.; Smith, A.D.; Box, T.W. 1975. Range management. New

York: McGraw-Hill, Inc. 532 p.

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temperatures as related to subsurface waterflows originating from

irrigation. Journal of Range Management. 51: 88–90.

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condition. Tech. Ref. 1737–9. Denver: Bureau of Land Management

Service Center. 51 p.

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condition and supporting science for lotic areas. Tech. Ref. 1737–15.

Denver: National Applied Resource Sciences Center. 127 p.

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of an open and a closed watershed. Journal of Environmental Health. 30:

157–163.

Chapter 14
158

Chapter 15

 Wildlife

Arthur R. Tiedemann1

Introduction

Numbers of large herbivores—mainly deer and elk—are

increasing throughout the United States (Riggs and others,

in press) with a commensurate increase in potential for them

to have an effect on the quality of drinking water sources. In

general, activities that provide healthy habitat usually help

maintain—or at least do not significantly degrade—water

quantity and quality for domestic use. However, contamination of surface waters by wild herbivores and diseaseorganism transmission among domestic herbivores and

wide-ranging large herbivores are major issues that must be

addressed by land managers. Emphasis of this chapter is the

role of large, wild herbivores, such as deer, elk, and moose

on microbiological quality of wildland surface water.

Information is also provided on the role of beavers and

muskrats.

Issues and Risks

Presence of disease-causing organisms in wildland surface

waters is the most critical aspect of problems related to the

influence of wildlife on water quality. There is no doubt that

the human pathogens Giardia spp. and Cryptosporidium

spp. (appendix D), and pathogenic Escherichia coli (E. coli)

are carried by and can be spread by a wide variety of

wildlife.

The impact of large, wild herbivores, such as deer, elk, and

moose, on water quality is an elusive and difficult problem.

These animals range widely and unpredictably, and their

densities and movements relative to surface waters are very

difficult to quantify.

Presence and numbers of fecal coliform (FC) in a given

volume of streamwater [usually number per 100 milliliters

(ml)] are used as indicators of the potential for presence of

disease organisms. At FC counts between 1 and 200 per 100

ml of water, the percentage occurrence of salmonella disease

organisms is 28 percent (Geldreich 1970). Occurrence of

salmonella increases to 85 percent for FC counts of 200 to

2,000 per 100 ml of streamwater and to 98 percent at FC

levels in excess of 2,000 per 100 ml. Bohn and Buckhouse

(1985) comprehensively reviewed the use of FC as an

indicator of wildland water quality. Use of FC counts as an

indicator of disease bacteria has several drawbacks. For

example, it is not a satisfactory indicator for Giardia.

Fecal streptococcus (FS) counts are also used as indicators

of the presence of disease organisms in water (Sinton and

others 1993), but there are no established standards as there

are with FC.

The ratio of FC to FS is one measure of the source of

bacterial contamination of wildland surface water

(Geldreich 1967). Geldreich (1976), Geldreich and Kenner

(1969), and Van Donsel and Geldreich (1971) established

ranges of FC to FS—in feces for humans, >4; cattle, 1.2 to

0.8; cattle and wildlife mixed, 0.08 to 0.04; and wildlife,

<0.04. Tiedemann and Higgins (1989) and Tiedemann and

others (1988) applied the concept to wildland watersheds

where several watersheds received only wildlife use and

others received varying degrees of cattle use and wildlife

use. For watersheds with no cattle use, the FC to FS ratio

was <0.04 for 90 percent of the samples collected. On

watersheds with intensive cattle use of 7 acres [2.8 hectares

(ha)] per animal unit month (aum), the FC to FS ratio in 75

percent of the samples was between 1.2 and 0.8. On

watersheds with lower intensities of cattle use [19 to 20

acres (7.7 to 8.1 ha) per aum], the ratio suggested a mixture

of cattle and wildlife use. Thus, the ratio established by

Geldreich appears to have potential as a tool to separate

wildlife and cattle sources of pollution. Baxter-Potter and

Gilliland (1988) also concluded that the ratio was useful for

distinguishing contaminant sources if the pollution had not

aged in the stream.

Recent efforts have focused on determination of the presence and sources of actual disease organisms in surface

water. Outbreaks of severe, bloody diarrhea associated with

E. coli [O antigen 157 and H antigen 7 (O157:H7)]

(Armstrong and others 1996) and numerous reports of

Giardia infection (Moore and others 1993) have been

1 Scientist Emeritus, USDA Forest Service, Pacific Northwest Research

Station, Wenatchee, WA.
159

responsible for emphasis on these organisms. Most cases of

E. coli O157:H7 infection and death have been caused by

consumption of infected meat (Armstrong and others 1996),

including deer jerky (Keene and others 1997). However,

multiple illnesses and deaths have occurred as a result of

drinking contaminated municipal water (Swerdlow and

others 1992). In the case of E. coli, until 1982, this organism

had not been considered a disease-causing entity (Riley and

others 1983, Wells and others 1983). As discussed previously, E. coli was formerly considered to be benign and an

indicator organism for the presence of organisms that

actually cause diseases.

Severe illness and deaths resulting from E. coli O157:H7

have prompted accelerated research to determine presence

and quantities occurring in various animal species including

wildlife. This is usually done by culture of samples of fresh

fecal material (Atwill and others 1997, Goodrich and others

1973, Monzingo and Hibler 1987), entrail samples from

harvested animals (Atwill and others 1997, Frost and others

1980), or samples from live animals (Erlandsen and others

1990). Transmission among domestic and wildlife species

(especially those with wide ranges) is a crucial facet of this

problem with far-reaching implications for drinking water

quality. Large herbivores, such as deer, elk, and moose, are

without boundaries and may serve as a vector for transmission of disease organisms among domestic livestock in

pastures or on rangelands. This concern is reinforced by

results of Sargeant and others (1999), who found similar

genetic strains of E. coli O157:H7 in cattle and free-ranging

white-tailed deer sharing the same pasture. Both types of

animals serve as reservoirs for E. coli O157:H7. Although

their conclusions were directed at consumption of deer

meat, their results suggest that we should be alert to potential problems with the quality of surface drinking water

sources.

Free-ranging white-tailed deer have also been shown to be

potential sources of contamination of Giardia and C.

parvum (Rickard and others 1999). These authors concluded

“the abundance of water throughout the State (Mississippi)

coupled with an overpopulation of white-tailed deer

indicates this cervid may pose a threat to surface water of

this area.”

Relations among wild herbivores and water quality are

highly complex involving the physical and vegetal nature of

the landscape, characteristics of surface water, the species of

wildlife, interrelations among wildlife and domestic

animals, and the disease or indicator organism involved.

Land management activities that tend to concentrate wildlife

or result in increased densities also will influence water

quality. Landscape characteristics that would tend to

influence the effect of wild herbivores on water quality

would be those that encourage or discourage high animal

concentrations near surface water, such as extremely steep

terrain or very dense vegetation. Stream characteristics that

influence organism concentrations are discharge, turbidity,

conductivity, pH, and temperature. Diurnal fluctuations of

bacterial counts are related to discharge, but evidence is

contradictory. Bohn and Buckhouse (1985) provide detailed

analyses of watershed characteristics that influence FC

numbers in streamwater. According to these authors,

bacterial counts generally increase with increasing discharge. Increased counts are apparently the result of

increased flushing of streambanks and associated fecal

deposits. Sediment transport and associated organisms may

also account for some of the increase in counts with

increasing discharge. Counts of FC have been shown to be

much higher in sediments than in surface water (Skinner and

others 1984, Stephenson and Rychert 1982). Counts of FC

seem to be inversely related to stream temperature

(Geldreich and Kenner 1969). Extremes of pH also influence FC viability with 5.5 to 7.5 as an optimal range

(McFeters and Stuart 1972). Type of wildlife species present

is a major determinant of the type and number of disease

organisms present in surface water. Wildlife closely associated with water, such as beavers and muskrats, may cause a

protracted, relatively stable contamination problem that may

be fairly predictable. Contamination by wider ranging

wildlife, such as deer and elk, will likely be transient,

sporadic, and somewhat unpredictable. Type of organism

present also depends to some degree on the species of

wildlife present. Giardia, for example, appears to be more

closely tied to presence of water-associated wildlife, such as

beavers and muskrats, than deer and elk (Monzingo and

Hibler 1987). This does not exclude deer and elk as an

element of the cycle of the organism or as a transmission

mechanism. The disease-causing organism E. coli O157:H7

appears to be more closely related to ruminants than to

animals with simple stomachs (Armstrong and others 1996).

Persistence of fecal organisms in surface water is an

important consideration. In the case of domestic animals,

FC levels in streamwater remain high for many months after

cattle are removed (Jawson and others 1982, Tiedemann and

others 1987). We don’t know if the same is true for wildlife

effects on water quality. Fecal coliform do survive for long

periods—at least 1 year in feces of both domestic (Clemm

1977) and wild herbivores (Goodrich and others 1973).

Complexity also arises from the rigorous sample collection,

transport, and analytical procedures that must be followed to

obtain accurate results. Isolation of actual disease organisms

is a complicated laboratory procedure that may also be very

costly. Sampling schedules must be carefully adhered to.

Chapter 15
160

Wildlife

Land treatment measures such as forest thinning, prescribed

burning, shrub control, seeding, and fertilization have the

potential to alter wild herbivore-use patterns. Some treatments, such as prescribed burning, are done to improve

wildlife habitat. These treatments have direct effects on

water quality that are discussed in chapters 10, 12, and 14 of

this report. Because the treatments have the potential to also

alter wild herbivore-use patterns, there is a potential for

indirect effects on water quality. If these habitat modifications concentrate animals near streams and near domestic

water supply withdrawals, water-quality impacts could be

serious. However, information is lacking on effects of wild

herbivores on water quality associated with changes in

habitat and resultant alterations in herbivore-use patterns.

Findings from Studies

Very little research has been conducted on large, wild

herbivores or beavers and muskrats as the identifiable

sources of fecal contamination of wildland surface water.

Results of studies of maximum levels of FC and Giardia

resulting from wild herbivore use are portrayed in table

15.1. Walter and Bottman (1967) provide some of the

earliest information on effects of wild herbivores on water

quality. They studied FC counts in two watersheds that serve

as a source of municipal water for Bozeman, MT. One

watershed was used by recreationists; the other was fenced

and patrolled to exclude the public and livestock. Thus, the

only potential source of FC in the closed watershed was

wildlife. Deer, elk, and moose were present in the closed

watershed in undetermined numbers. They found maximum

FC counts exceeding 200 per 100 ml of streamwater in the

closed watershed. Fecal coliform counts were actually much

higher in streamwater in the closed than in the open watershed, suggesting higher wildlife populations in the closed

watershed—perhaps as a result of high levels of human use

of the open watershed or watershed characteristics.

In another early study, Kunkle and Meiman (1967) measured maximum FC counts of 25 per 100 ml in streamwater

from an area essentially free from human impact—very

limited hiking or camping, and no domestic livestock

grazing. Wildlife species were not listed, but it is safe to

assume that the area is inhabited by mule deer.

Maximum FC (100+ per milliliter) and Giardia (1 per 100

ml) levels were also high in streamflow from a pristine,

forested watershed in western Washington. These levels

were presumably from wildlife contamination because

human use was very light [5 to 30 days per year per

0.6 miles (1 kilometer) of stream]. Deer, elk, beavers,

mountain beavers, river otters, and marmots were listed as

possible sources of contamination.

Streamwater from pristine land in southern Finland was

shown to contain high levels of FC (maximum, 268 per 100

ml of streamwater). The high counts apparently resulted

from wild deer and moose use of the watershed (Niemi and

Niemi 1991).

In eastern Oregon, forested watersheds that supported no

domestic grazing had maximum FC counts in excess of 500

per 100 ml of streamwater (Tiedemann and others 1987).

Deer and elk were the predominant large herbivores

inhabiting the watersheds.

Maximum Giardia count observed by Monzingo and Hibler

(1987) was 0.06 per 100 ml of streamwater in beaver ponds

in Colorado.

Streamflow from watersheds without domestic livestock in

Utah (Doty and Hookano 1974) and Wyoming (Skinner and

others 1974) contained relatively high counts of FC—22 to

183 per 100 ml. Although FC origin was attributed to

wildlife in both studies, there was no mention of which

wildlife species were responsible. Recreational use in the

two studies was limited to hiking.

Reliability and Limitation of Findings

Research to date lacks replication or detailed examination of

species responsible, numbers of each species, and their

distribution. Its results, however, indicate that wild herbivores pose a risk to drinking water quality.

Studies spanned a relatively broad geographical scope and

suggest that drinking water quality may be a problem

wherever there is contact between wild herbivores and

surface water.

Changes in habitat, whether deliberate or uncontrolled, that

attract wildlife to surface drinking water sources may

increase the risk of introducing contaminants. Management

that discourages animals from concentrating near streams

that are sources of public drinking water may help alleviate

potential contamination problems.
161

Table 15.1—Streamwater fecal coliform and Giardia responses to herbivore use

Maximum counts observeda

Location Setting Herbivore Fecal coliform Giardia spp.b Reference

Colorado Beaver ponds Beaver Not measured 0.021 Monzingo and Hibler 1987

Washington Pristine forested Beaver, 100+ 1 Ongerth and others 1995

 watershed, low- mountain

 level human use beaver, otter,

 marmot

Southern

 Finland Pristine lands Elk, deer 268 Niemi and Niemi 1991

Eastern Forested Deer, elk 500+ Tiedemann and others 1987

 Oregon watersheds, no

 domestic livestock

 use

Montana Municipal Deer, elk 200+ Walter and Bottman 1967

 watershed, closed

 to public access

Colorado Area essentially Not indicated 25 Kunkle and Meiman 1967

 free of human

 impact, no domestic

 livestock use

Utah Watersheds with Not indicated 36–183 Doty and Hookano 1974

 no livestock use

Wyoming Watershed open Not indicated 22 Skinner and others 1974

 to hikers, no

 livestock use

a

Counts per 100 milliliters of streamwater.

b Locations with no data in this column indicate that they were not measured.

Research Needs

1. Carefully controlled experiments are needed to determine

the magnitude of the effect of wild herbivores on microbiological water quality. Such research must separate

effects of domestic herbivores and humans from those

that result from wildlife. Research must also document

the species of herbivore, the amount of time that each

spends near surface water, and organisms associated with

fecal deposits. Carefully designed water sampling

procedures must accompany these studies. Samples

above, within, and below areas of activity of sufficient

frequency to represent all hydrograph stages will be

essential.

2. Landscape-scale satellite animal tracking experiments

with large, wild herbivores, such as the Starkey Experimental Forest and Range studies in northeast Oregon

(U.S. Department of Agriculture, Forest Service, Pacific

Northwest Research Station, La Grande, OR), provide

opportunities to relate large herbivore movements by

species to changes in water quality. Rationale and

methodology of those studies could be used as a model

for design of water-quality studies. Starkey’s studies

(Rowland and others 1997) could also serve as a model

for research to determine relations among forest and

rangeland treatments, wild herbivore-use patterns, and

resultant changes in water quality.

Chapter 15
162

Wildlife

3. Perhaps it would be appropriate to explore advanced

technology to discourage large, wild herbivores, such as

the technique described by Tiedemann and others (1999)

for cattle. This technique would probably be appropriate

primarily where wild herbivore contamination of surface

water creates severe risks from drinking the water.

4. The literature shows that microbiological and epidemiological research on Giardia and E. coli O157:H7 is well

underway. Hancock and others (1998), however, pose

some problems that are urgently in need of resolution; E.

coli O157:H7 may: (1) have multiple species in which it

resides for long periods; (2) be capable of transiently

colonizing many species; and (3) have an environmental

source, such as sediments where it may be able to

multiply prolifically. Similar questions should be raised

with regard to Giardia.

Key Points

1. Presence of wild herbivores has the potential to influence

microbiological quality of wildland surface water and to

render it unsafe for drinking unless it is adequately

purified. It is important for wildland managers to understand the potential for a problem to exist even though

they cannot document the actual source of contamination.

2. Relationships between wild herbivores and water quality

are highly complex involving physical and vegetal

landscape characteristics, hydrologic parameters (discharge, pH, temperature), type of contaminating organisms, species and numbers of animals, use patterns of

animals, mixing of domestic and wild herbivores, length

of time since animals were present, and land treatment

measures that have been implemented.

3. Potential for presence of disease organisms such as

pathogenic E. coli O157:H7 and Giardia emphasizes the

urgency to understand the relations between wild

herbivores and water quality. No studies have related

water quality to wild herbivore species, numbers, and

land-use patterns.

4. Land treatment measures such as thinning, prescribed

burning, seeding, and fertilization, whether designed

specifically for wildlife habitat alteration or other

management purposes may alter herbivore-use patterns

and exert secondary effects on water quality. These

effects are presently undocumented and poorly understood.

5. Free-ranging large, wild herbivores may be vectors for

transmission of disease organisms among domestic

livestock. Associated contamination of surface water and

potential public health hazards create an urgency to

understand these relations and effects on water quality.

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A.R., eds. Managing interior Northwest rangelands: the Oregon range

evaluation project. Gen. Tech. Rep. PNW–238. Portland, OR: U.S.

Department of Agriculture, Forest Service, Pacific Northwest Research

Station: 57–91.

Tiedemann, A.R.; Higgins, D.A.; Quigley, T.M. [and others]. 1987.

Responses of fecal coliform in streamwater to four grazing strategies.

Journal of Range Management. 40: 22–329.

Tiedemann, A.R.; Higgins, D.A.; Quigley, T.M. [and others]. 1988.

Bacterial water quality responses to four grazing strategies—comparisons with Oregon standards. Journal of Environmental Quality. 17:

492–498.

Tiedemann, A.R.; Quigley, T.M.; White, L.D. [and others]. 1999.

Electronic (fenceless) control of livestock. Res. Pap. PNW–510.

Portland, OR: U.S. Department of Agriculture, Forest Service, Pacific

Northwest Research Station. 23 p.

Van Donsel, D.J.; Geldreich, E.E. 1971. Relationships of Salmonellae to

fecal coliforms in bottom sediments. Water Resources. 5: 1079–1087.

Walter, William G.; Bottman, Robert P. 1967. Microbiological and

chemical studies of an open and closed watershed. Journal of Environmental Health. 30: 157–163.

Wells, J.G.; Davis, B.R.; Wachsmuth, I.K. [and others]. 1983.

Laboratory investigation of hemorrhagic colitis outbreaks associated

with a rare Escherichia coli serotype. Journal of Clinical Microbiology.

18: 512–520.

Chapter 15
164

Chapter 16

Water Birds

Christopher A. Nadareski1

Introduction

This chapter synthesizes the available information on the

sources, mechanisms, and risks of drinking water contamination from water birds. By water birds, we mean geese,

ducks, and other duck-like swimming birds such as cormorants, gulls, and wading birds. In general, activities that

provide healthy habitat usually help maintain—or at least do

not significantly degrade—water quantity and quality for

domestic use. However, water birds have been implicated in

the contamination of both large and small drinking water

supplies.

Issues and Risks

The U.S. Department of Agriculture, Wildlife Services,

collect and disseminate information about nuisance-related

problems caused by a variety of birds. Birds commonly

create human health concerns by polluting potable drinking

water supplies and fouling recreation areas, such as swimming beaches. Species like the cormorant are thought to

negatively impact recreational fisheries and drinking water

supplies. Some terrestrial bird species, such as swallows,

pigeons, and starlings, may nest, feed, or roost in water

intake structures where their defecation has negative

impacts on water quality. Biological, chemical, and nutrient

pollutants may be released or deposited by water birds. The

type and quantity of pollutant will determine what kind and

level of water bird management is required to prevent

contamination to a reservoir system.

Estimating impacts from the various types of contamination

requires information on the species of water bird inhabiting

the water supply, the number of birds per species, the daily

defecation or pellet regurgitation rates, the amount of

contamination per specified weight of feces or pellet, and

the daily amount of activity by each bird on the water. The

principal pathogens of concern are listed in table 16.1.

Findings from Studies

Studies have found that water birds and some terrestrial

species affect the biological, nutrient, and chemical quality

of water.

Contamination

Biological

Among the various types of water bird contaminants are

excrement containing bacteria, protozoans, or enteric

viruses. These contaminants are routinely monitored by

water suppliers and regulated by the U.S. Environmental

Protection Agency (EPA) and various State health departments. Numerous studies have documented the occurrence

of fecal coliform bacteria and other pathogens in many

North American water bird species (Ashendorf and others

1997, Gould and Fletcher 1978, Hatch 1996, Hussong and

others 1979). As a result, bird control programs have been

developed and implemented at many municipal water

sources (Ashendorf and others 1997, Blokpoel and Tessier

1984). Water bird excrement has been reported to contain

both human and nonhuman pathogens. Microbiological

analysis for human bacterial pathogens found in drinking

water are generally represented by the fecal coliform

bacteria group which is used as an index to identify the

probability one or more of these organisms will be present

in the sample (chapter 2 and table 16.1). Fecal coliform

originates solely from warm-blooded animals , and analytical methods to identify the origin of bacterial type (human

versus nonhuman) are currently under investigation. These

new methods require an expensive and time-consuming

examination of the sample. As a result, most public water

suppliers, at best, can only speculate on the origin of the

drinking water bacterial contamination. In the larger

municipal water systems, such as New York City’s, methods

are being developed to link bacterial sources from water

samples to human, nonhuman, or both sources.

1

 Research Scientist, New York City Department of Environmental

Protection, Valhalla, NY.
165

Table 16.1—Indicator and principal pathogens of concern in contaminated drinking water

Organism type Disease Symptoms

Fecal coliform bacteria group

Bacteria

Escherichia coli, Klebsiella Diarrhea, dysentery, Diarrhea, nausea, cramps, fever, vomiting, pneumonae hemorrhagic colitis mucus in stools

Important human pathogenic

organisms suspected from feces

Bacteria

Shigella spp. Shigellosis Diarrhea, fever, cramps, tenesmus, blood in

 stools

Salmonella typhimurium Salmonellosis Abdominal pain, diarrhea, nausea, vomiting,

 fever

S. typhi Typhoid fever Abdominal pain, fever, chills, diarrhea or

 constipation, intestinal hemorrhage

Enterotoxigenic Diarrhea Diarrhea, fever, vomiting

Campylobacter jejuni Gastroenteritis Abdominal pain suggesting acute appendicitis,

 fever, headache, malaise, diarrhea, vomiting

Vibrio cholerae (incidence rare

or negligible in the United

States) Gastroenteritis Vomiting, diarrhea, dehydration

Yersinia spp. Plague, hemorrhagic Diarrhea, fever, abdominal pain

 enterocolitis, terminal

 ileitis, mesenteric

 lymphadenitis,

 septicemia

Enteric viruses

Hepatitis A virus Hepatitis Fever, malaise, anorexia, jaundice

Norwalk-like agent Gastroenteritis Diarrhea, abdominal cramps, headache, fever,

 vomiting

Virus-like 27 nm particles Gastroenteritis Vomiting, diarrhea, fever

Rotavirus Gastroenteritis Vomiting followed by diarrhea for 3 to 8 days

Protozoa

Giardia lamblia Giardiasis Chronic diarrhea, abdominal cramps,

 flatulence, malodorous stools, fatigue, weight

 loss

Cryptosporidium Cryptosporidiosis Abdominal pain, anorexia, watery diarrhea,

 weight loss; immuno-compromised individuals

 may develop chronic diarrhea

Entamoeba histolytica Amebiasis Vary from mild diarrhea with blood and

 mucus to acute or fulminating dysentery with

 fever and chills

nm particles = nanometers.

Source: Murray and others 1995, Hurst and others 1997.

Chapter 16
166

Water Birds

Nutrients

Contributions of nutrients to aquatic ecosystems by water

bird excrement have been well documented (Gould and

Fletcher 1978, Manny 1994), but debate continues over the

significance of impacts from bird defecation on nutrient

loading (Hoyer and Canfield 1994, Murphy 1984). Among

the various nutrients identified in water bird excrement,

nitrogen and phosphorus are of greatest interest. Both have

the potential to increase the rate of eutrophication and

degrade the quality of a drinking water supply

(Vollenweider and Kerekes 1980).

Chemical

Another type of water bird contamination may be of

chemical origin. Glandular releases, body oils, pesticides,

and other hydrocarbons are transported by the birds. The

impact to drinking water supplies from glandular releases

and body oils emitted by water birds is not well understood.

Chemical pollutants carried externally on feathers or

vectored by water birds, although not well documented, may

potentially cause contamination of a water supply. For

example, many reservoir systems do not provide adequate

food supplies for gulls or other opportunistic feeders. Gulls,

therefore, will seek alternative foraging locations, such as

agricultural areas, urban centers, moist fields, and landfills,

where they may be exposed to a variety of chemicals either

through ingestion or external attachment. The gulls may

accumulate various chemicals on their feathers and transport

them back to the reservoir where they roost. Gulls may also

ingest contaminated materials from landfills or sewage

treatment facilities, carry them back to the reservoir, and

regurgitate them. Additional studies are needed to determine

these impacts.

Water Birds as Vectors of Contamination

A variety of human-related activities, such as urbanization,

resource exploitation, agriculture, and land conservation,

have the potential to promote or discourage populations of a

variety of water bird species.

Water birds deposit pollutants in drinking water supplies

during roosting, foraging, and overflights. The birds may

acquire pathogens, such as Giardia spp. or Cryptosporidium

spp., from domestic farm animals (Graczyk and others

1998) and from urban centers such as shopping malls or

landfills. Agricultural operations, such as the spreading of

fodder or manure, growing crops, and tilling soil, offer

attractive foraging locations for some species that travel

great distances. In a study of Canada geese foraging on

agricultural land near the Chesapeake Bay, researchers

identified a high incidence of Giardia and Cryptosporidium

in their fecal matter (Graczyk 1996) compared to an

extremely low incidence of the same two protozoans in

geese sampled at an urban reservoir 15 miles north of New

York City.

The pathways by which water contaminated with biological

materials, chemicals, and nutrients can enter into the

reservoir from water birds include streamflow through

drainage basins, stormwater surface sheet flow, fly-over

fecal releases, and direct fecal deposition. The location of

the source of contamination with respect to a reservoir will

determine the degree of water-quality impacts. Size and

water flow patterns in relation to the water intake structure

from which drinking water is drawn also affects the extent

of impacts. The extent of contamination depends on the dieoff rate, settling rate, and water travel time.

Seasonality of Impacts

Pollutant loadings from water birds vary widely by season.

Migratory birds may cause problems only during brief

stopovers. Local breeding populations of water birds may

also have negative impacts, and those impacts may persist

through much of the year. The location of bird activity

(roosting, breeding, foraging, etc.) on the reservoir relative

to a water intake facility may also determine importance of

strategies for managing water bird populations.

Reliability and Limitations of Findings

The types of water-quality impacts are understood. The

magnitudes of impacts are less well understood. Resident

and migratory water birds will pose similar problems to

water supply reservoirs. The magnitude of the impact will

vary by species and number of birds inhabiting the reservoir,

the surrounding environs and associated sources of contamination, and seasonal patterns of migration. Therefore, water

bird impacts will vary and need to be assessed for each

reservoir.
167

Key Points

1. Land managers may have different objectives concerning

habitat management for water bird species than reservoir

managers. Both forest managers and reservoir managers

can and do influence the population dynamics of water

bird species. Their management objectives can conflict;

for example, the forest manager may want to increase

populations of certain water bird species while the

reservoir manager may try to decrease or eliminate their

activity on or near the reservoir (see New York City case

study).

2. To address the contamination of drinking water supplies

by water birds, a comprehensive watershed protection

plan should be drawn up. This plan needs to identify all

major sources of pollution. If it is determined that water

birds are major contributors, an additional plan for water

bird management should be developed. This water bird

management plan should also incorporate objectives both

for the land manager and the reservoir manager. Populations of all water bird species that breed or migrate

throughout the watershed should be inventoried to

identify all potential impacts on water quality. The type

of pollution should be identified, as should its potential

origins. Where the source of the contaminant is human

activity and birds are only a vector, it may be possible to

control the contamination by changing the offending

human activity.

Case Study: New York City Waterfowl

Management Program

As the Nation’s largest, unfiltered water supplier, the city of

New York Department of Environmental Protection (DEP)

is responsible for the maintenance of 19 reservoirs, 3

controlled lakes encompassing an area of almost 2,000

square miles, and serving 9 million consumers. Beginning in

the early 1980’s, the city enhanced its water-quality monitoring programs to address watershed protection issues. As

more stringent Federal regulations were implemented

through the Safe Drinking Water Act of 1986 and subsequent Surface Water Treatment Rule of 1989, New York

City remained determined to maintain its unfiltered status

through stringent criteria set by the EPA. In order to fulfill

this filtration avoidance determination, the DEP developed a

comprehensive Watershed Protection Program (WPP) to

identify and eliminate sources of pollution that would

compromise its water quality. An important component of

the WPP was the implementation of a Waterfowl Management Program that identified birds, particularly gulls, geese,

ducks, and cormorants, as the primary source of fecal

coliform bacteria to the water system. Baseline bird

population data was well correlated with the seasonal

elevations of bacteria. As a result, DEP instituted a bird

deterrent/harassment program to eliminate the presence of

both breeding and migratory Canada geese and migratory

gulls, ducks, and other water birds. Techniques used to deter

the birds included shoreline fencing, meadow management

to deter feeding opportunities, bird distress tapes, and the

use of pyrotechnics from motorboats and hovercraft from

dawn to dusk on a daily basis. The techniques were highly

effective: elimination of defecation from roosting birds

eliminated the seasonal elevations of bacteria, allowing New

York City to maintain its filtration avoidance status.

Literature Cited

Ashendorff, A.; Principe, M.A.; Seeley, A. [and others]. 1997. Watershed

protection for New York City’s supply. Journal of the American Water

Works Association. 89(3): 75–88.

Blokpoel, G.; Tessier, G.D. 1984. Overhead wires and monofilament lines

exclude ring-billed gulls from public places. Wildlife Society Bulletin.

12: 55–58.

Gould, D.J.; Fletcher, M.R. 1978. Gull droppings and their effects on

water quality. Water Research. 12: 665–672.

Graczyk, T.K.; Fayer, R.; Trout, J.M. [and others]. 1998. Giardia sp.

cysts and infectious Cryptosporidium parvum oocysts in the feces of

migratory Canada geese (Branta canadensis). Applied and Environmental Microbiology. 64(7): 2736–2738.

Hatch, Jeremy J. 1996. Threats to public health from gulls (Laridae).

International Journal of Environmental Health Research. 6: 5–16.

Hoyer M.V.; Canfield, D.E. 1994. Bird abundance and species richness on

Florida lakes—influence of trophic status, lake morphology, and aquatic

macrophytes. Hydrobiologia. 280: 107–120.

Hurst, C.J.; Knudsen, G.R.; McIvering, M.J. [and others]. 1997.

Manual of environmental microbiology. Washington, DC: American

Society for Microbiology Press. 894 p.

Hussong, D.; Damare, J.M.; Limpert, R.J. [and others]. 1979. Microbial

impact of Canada geese (Branta canadensis) and whistling swans

(Cygnus columbianus columbianus) on aquatic ecosystems. Applied and

Environmental Microbiology. 37(1): 14–20.

Manny, B.A.; Johnson, W.C.; Wetzel, R.G. 1994. Nutrient additives by

waterfowl to lakes and reservoirs: predicting their effects on productivity

and water quality. Hydrobiologia. 279: 121–132.

Murphy, S.M.; Kessel, B.; Vining, L.J. 1984. Waterfowl populations and

limnological characteristics of taiga ponds. Journal of Wildlife

Management. 48(4): 1156–1163.

Murray, P.R.; Baron, E.J.; Pfaller, M.A. [and others]. 1995. Manual of

clinical microbiology. 6th ed. Washington, DC: American Society for

Microbiology Press. 1,482 p.

Chapter 16
168

New York City Department of Environmental Protection. 1993.

Implications of phosphorus loading for water quality in NYC Reservoirs.

Corona, NY: New York City Department of Environmental Protection.

65 p.

U.S. EPA. 1999. U.S. EPA’s unfiltered water supply systems guidance

manual—revised draft. Washington, DC: U.S. Environmental Protection

Agency. 58 p.

Vollenweider, R.A.; Kerekes, J.J. 1980. Background and summary results

of the OECD cooperative program on eutrophication. EPA 440/5–81–

010:25–36. Washington, DC: Environmental Protection Agency: 25–35.

Water Birds
169

Chapter 17

Fish and Aquatic Organisms

C. Andrew Dolloff 1

Introduction

Freshwater fish management can impact water quality by

manipulating fish and other organisms and by altering the

physical, chemical, or biological attributes of habitat.

Fishery managers manipulate populations for a variety of

reasons, including attempts to maximize yield, control size

and age structure, establish populations of desirable and

remove populations of undesirable species, and to reestablish populations and communities that have been extirpated.

Habitat management often accompanies attempts to manipulate populations. All components of habitat are subject to

manipulation, including not only the obvious structural

elements, such as substrate, cover, and flow obstructions,

but also water chemistry and the content of the biological

community.

Fish Hatcheries and Aquaculture Facilities

Issues and Risks

Fish hatcheries or fish culture facilities can reduce water

quality. Untreated effluent from these facilities typically

consists of metabolic waste products and solids derived

from uneaten fish food and fish wastes (Goldburg and

Triplett 1997). In addition, culture facilities also may

intermittently discharge pathogenic bacteria and parasites

and the chemicals and drugs used in the prevention or

treatment of disease (Liao 1970, Piper and others 1986). The

usual waste stream from a culture facility can be treated as a

chronic point source and effluent quality and quantity can be

monitored at or immediately below the outfall. Monitoring

results tend to be highly variable because of variation in

production schedules and other activities (Foy and Rosell

1991) such as periodic cleaning and flushing (Bergheim and

others 1984).

Findings from Studies

The quality of effluent water varies considerably depending

on the specific features of the culture system including the

species of fish, intensity of production, the diet and feeding

regime, and the temperature and chemical character of

source water (Axler and others 1997). The four most

important dissolved constituents of fish cultural wastewater

include ammonia, nitrate, phosphate, and organic matter.

Ammonia is toxic to most aquatic life and all four constituents are primary agents of eutrophication. Along with

dissolved organic matter, suspended solids, which are

predominantly organic, contribute to biological oxygen

demand in receiving waters. Nutrient loading can be

significant, as 60 to 75 percent of the nitrogen and phosphorus in fish food ultimately becomes part of the waste stream

(Axler and others 1997). Unless intakes for domestic water

supplies are located immediately downstream from the

outfall, however, hatchery effluents are not likely to severely

degrade domestic water supplies.

Effluents from trout farms may diminish water quality

slightly during periods of low flow and high temperature

(Selong and Helfrich 1998). In a study of five Virginia trout

farms, Selong and Helfrich (1998) found that total ammonia

nitrogen, un-ionized ammonia nitrogen, and nitrite nitrogen

levels increased downstream from effluent outfalls but did

not exceed thresholds for lethal exposure for aquatic

organisms. Dissolved oxygen levels also decreased but were

typically above 7.0 parts per million (ppm). Temperature,

pH, nitrate nitrogen, and total phosphorus levels did not

differ from upstream levels. Substrate embeddedness was

greater below outfalls from two farms but settleable solids

concentrations were always <0.1 ppm. The lack of significant water-quality degradation reflected the tendency of

growers to adjust production to correspond to periods of

high flow and low-to-moderate temperature (Selong and

Helfrich 1998). In Washington State, however, hatchery

effluents during the summer had significantly elevated

temperature, pH, suspended solids, ammonia, organic

nitrogen, total phosphorus, and biological oxygen demand

(Kendra 1991). Phosphorus loading in hatchery effluent in

one Washington State creek was equivalent to secondarily

treated sewage discharge from a town of 2,300 people. The

1 Project Leader, USDA Forest Service, Southern Research Station,

Blacksburg, VA.
170

influence of hatchery effluents tends to be localized in the

immediate downstream reach (Doughty and McPhail 1995).

Effects on periphyton and macroinvertebrates were not

detected about 1,200 feet [366 meters (m)] downstream of

the outfalls from five Virginia trout farms (Selong and

Helfrich 1998).

Because most of the total nutrient load produced at a typical

facility is in the form of settleable solids, treatment usually

consists of diverting wastewater through settling basins

before discharge (Piper and others 1986). Much of the

variation observed in effluent waste among facilities can be

traced to differences in settling characteristics (Mudrak

1981). When located near towns or cities, culture facilities

can avoid discharging directly into streams or rivers by

connecting to municipal sewage treatment facilities.

Fish disease organisms, chemicals, and other additives

commonly used in culture facilities also may influence the

quality of water for domestic use. Fish trematodes, cestodes,

and nematodes can be transmitted to people who eat certain

species of raw fish. In general, however, fish diseases do not

present risks to human health (Hoffman 1999). Most

potential problems with fish disease and water quality arise

from the chemicals and procedures used in treatment. At

present, relatively few chemicals have been approved by the

U.S. Food and Drug Administration (FDA) for use in the

treatment of fish diseases. The antibacterials approved for

use on food fishes include oxytetracycline, sulfamethoxine

(Romet), and sulfamerazine. Most drugs are administered

through feed, from which a considerable fraction may be

released into the environment. Release occurs through three

routes: uningested food, feces, and in urine and bile fluid

(Capone and others 1996). Residues of oxytetracycline in

sediments under net pens of intensively cultured (and

treated) Atlantic salmon were present for at least 10 months

after treatment (Capone and others 1996). Other drugs

approved for use in treating fish include formalin (for the

treatment of external parasites and fungus on fish eggs) and

tricaine methanesulfonate, an anesthetic. The FDA also

maintains a list of unapproved new animal drugs (drugs that

have not gone through the formal approval process but are

not expected to have a negative impact on the environment)

for use in aquaculture of food fishes provided the following

conditions are met:

1. The drugs are used for the prescribed indications,

including species and life stage;

2. The drugs are used at the prescribed dosages;

3. The drugs are used according to good management

practices;

4. The product is of an appropriate grade for use in food

animals; and

5. An adverse effect on the environment is unlikely.

Among a host of common compounds and substances in this

classification are sodium chloride and ice. Examples of

unapproved drugs for use in aquaculture are acetic acid

(parasiticide for fish), calcium chloride (ensures proper egg

hardening, aids in maintaining osmotic balance in fish),

carbon dioxide gas (fish anesthetic), hydrogen peroxide

(fungicide), magnesium sulfate (controls monogenetic

trematode and external crustacean infestations), potassium

chloride (relieves osmoregulatory stress and prevents

shock), povidone iodine compounds (fish egg disinfectant),

sodium sulfate (improves egg hatchability), and urea and

tannic acid (denatures adhesive component of fish eggs).

In addition to treating specific diseases, all culture facilities

must undergo periodic cleaning and sterilization. Chlorine

(HTH) is often used for this purpose; exposure to 5 ppm for

1 hour kills nearly everything. Chlorine rapidly loses its

toxicity (1 day or less) and can be neutralized by sodium

thiosulfate.

Reliability and Limitations of Findings

A large body of research has accumulated on the design,

construction, operation, and maintenance of fish culture

facilities. Because a significant portion of this research has

addressed the issue of effluent management, the findings are

considered highly reliable. These findings are widely

applicable to flowing water.

Research Needs

None identified.

Key Points

Existing knowledge will permit managers and policymakers

to make informed decisions about the impact of culture

facilities on domestic water supplies.

Fish and Aquatic Organisms
171

Chemical Reclamation

Issues and Risks

Piscicidal (fish-killing) compounds have been used to

sample, control, or eradicate fish populations since the

1930’s (Bettoli and Maceina 1996). Purposes have included

eradication of exotic species such as common carp from

ponds, lakes, and reservoirs and removal of nonnative

species from headwater streams. With the exception of

sampling programs, the objective of most piscicide applications is to remove one or more species so that the water

body can be stocked soon after poisoning with species

considered more desirable.

Findings from Studies

Only registered piscicides may be applied to water in North

America. Among a variety of possible compounds, only

rotenone (trade name Noxfish), antimycin A (trade name

Fintrol), 3-trifluoromethyl-4-nitrophenol (TFM), and

Bayluscide (a nitosalicylanilide salt, trade name Bayer 73)

are approved for use in the United States. Bayluscide and

TFM, either alone or in combination, are used exclusively to

sample or control sea lamprey, primarily in streams flowing

into the Great Lakes (Bettoli and Maceina 1996, Marking

1992). In the United States, lampricides can be applied only

by personnel certified by the U.S. Fish and Wildlife Service

or an approved State conservation agency (Bettoli and

Maceina 1996).

Rotenone, a naturally occurring chemical derived from

the roots of tropical plants in the genera Derris and

Lonchocarpus, has been used for centuries by native people

to kill fish for food. Rotenone is highly toxic to fish, which

are killed by disruptions in cellular respiration (Haley 1978).

Most nontarget aquatic organisms usually are not affected

by the concentrations of rotenone used to kill fish, but high

doses can kill amphibians, some reptile species, and a

variety of macroinvertebrates (Bettoli and Maceina 1996).

The toxicity and persistence of rotenone are influenced by

turbidity, temperature, and pH; in general, rotenone is most

toxic in clear, warm, acidic water. Residues of rotenone

were detectable in the bottom sediments of experimental

ponds in Wisconsin for nearly 14 days during the spring at

46 o

F (8 oC), but decreased below the limits of detection

within 3 days during the summer at 72 oF (22 oC) and during

the fall at 59 oF (15 oC) (Dawson and others 1991). Higher

temperatures coupled with the presence of clay (bentonite)

led to adsorption and rapid decline in residues. Dawson and

others (1991) suggested that at warmer temperatures, water

treated with rotenone would be safe for swimming immediately after treatment with a concentration of 250 parts per

billion (ppb). A strong oxidizing agent such as potassium

permanganate (applied at 2.0 to 2.5 ppm) will detoxify

rotenone when applied downstream of treatment sections in

flowing water or in lakes where managers wish to limit the

area of kill.

Rotenone has a low mammalian toxicity (Marking 1988)

and relatively short half-life (<1 day in water at 73 o

F)

(Gilderhus and others 1988). Nevertheless, negative public

perceptions about the use of any poisonous substance in

open water have curtailed the use of rotenone over the last

decade. Older formulations of rotenone, no longer approved

for sale in the United States, sometimes contained carrier

substances such as trichloroethylene and piperonyl butoxide,

which are known carcinogens. Residues of the latter were

detected up to 9 months after treatment of Lake Davis in

California and caused disruption of a public drinking water

supply that drew national publicity.

Similar to rotenone, antimycin is relatively nontoxic to

mammals (Herr and others 1967) but highly poisonous to

fish. Antimycin toxicity in water varies by species, and

concentrations over 100 ppb may be required to kill resistant

species (Bettoli and Maceina 1996). The effectiveness of

antimycin may decrease below 50 o

F (10 oC) (Tiffan and

Bergersen 1996), but in warm water, temperature does not

greatly influence its toxicity. High turbidity (Gilderhus

1982) and alkalinity (Tiffan and Bergersen 1996) decrease

its effectiveness and persistence (Bettoli and Maceina 1996,

Lee and others 1971). Turbulence leading to oxidation and

foaming also may contribute to antimycin inactivation

(Tiffan and Bergersen 1996).

Reliability and Limitations of Findings

A large body of research has accumulated on the short- and

long-term effects of chemical reclamation. Studies have

been conducted in both warmwater and coldwater habitats

under widely varying conditions. The findings are considered highly reliable and widely applicable.

Research Needs

None identified.

Key Points

Existing knowledge will permit managers and policymakers

to estimate impacts of piscicides on fish populations and

water quality. However, controversy likely will continue

over the use of piscicides in sources of public water

supplies.

Chapter 17
172

Restoration and Reintroduction

of Populations and Communities

Issues and Risks

With the growth in environmental awareness in recent years,

management agencies increasingly are called upon to restore

species that have been intentionally or accidentally extirpated from aquatic ecosystems. In most situations, the

impact of the restoration on water quality will be almost

entirely positive because improved water quality is one of

the prerequisites for survival of target organisms. Managers

must be aware, however, that one of the most obvious signs

of a successful restoration—large numbers of fish—may

contribute to periodic, temporary declines in water quality.

Large numbers of fish may cause problems with water

quality during periods of low flow and high temperature if

the fish become stressed, die, and decompose. Large

numbers of fish of a single species seldom die in a small

area under such conditions except when anadromous fish

such as Pacific salmon or alosids (shads and alewifes)

congregate for spawning.

Findings from Studies

Although some aspects of water quality may be temporarily

degraded by the decay of anadromous fish carcasses, the

long-term effect can have great ecological value. Nutrients

from anadromous fish carcasses help maintain the productive capacity of streams and riparian zones in coastal

watersheds of the Pacific Northwest and Alaska (Cederholm

and others 1989, 1999; Wilson and Halupka 1995) and

streams of the Atlantic Coastal Plain (Garman and Macko

1998). On the west coast, the carcasses of abundant pink

(Brickell and Goering 1970) and sockeye salmon (Kline and

others 1994) and the less abundant coho salmon (Bilby and

others 1996) contribute important nutrients, particularly

nitrogen, to otherwise nutrient-poor watersheds. Similar

relationships have been observed in Atlantic slope drainages

(Durbin and others 1979, Garman 1992).

Accidental introductions generally are of greater concern

both ecologically and for water quality. Perhaps the most

troublesome, accidentally introduced aquatic pest is the

zebra mussel Dreissena polymorpha. This bivalve mollusk

first arrived in North America in ship’s ballast via the St.

Lawrence Seaway sometime around 1986 (Hebert and

others 1989). Since that time, it has invaded all of the Great

Lakes and the major rivers of the Eastern United States

(Ludyanskiy and others 1993). Invasion of virtually all the

major river systems in North America is viewed by some as

inevitable, with only the specific timetable subject to

question (Morton 1997). Zebra mussels will attach to

virtually any surface, including the interior of water inlet

pipes. Zebra mussels also will attach to the hulls of both

large, commercial and smaller recreational vessels, which

act as dispersal agents (Keevin and others 1992). The

primary economic cost of the zebra mussel invasion has

been the fouling of intakes for raw industrial and potable

water. Zebra mussels may be present in numbers sufficient

to clog water intakes, necessitating either abandonment or

laborious and repeated cleaning of the pipes. Over $4 billion

per year may ultimately be spent on attempts to control or

mitigate zebra mussel impacts (Morton 1997). The ecological impacts of the invasion have been equally profound,

particularly on native unionid mussels (Haag and others

1993). Nalepa and Schloesser (1993) completely reviewed

issues surrounding the zebra mussel invasion.

The accidental introduction of the zebra mussel, while

certainly regrettable, may have future benefits. Morton

(1997) noted that in Europe, where the species has been

established for around 170 years, zebra mussels are used as

biomonitors for trace metals and radionuclides. Their natural

water filtering abilities have made them useful both for

restoration of natural water systems and in the treatment of

human sewage. With development of genetically sterile

stocks, it may be possible to employ zebra mussels or other

suitable species in the cleanup of both natural and artificial

water bodies in North America (Morton 1997).

Reliability and Limitations of Findings

Although this area of research and restoration is relatively

recent, the findings are generally considered reliable.

Restoration ecology is one of the newest branches of

ecology. Relatively little research has accumulated and there

are few long-term studies of the effects of reintroduced or

accidentally introduced species. The likely impacts of exotic

invasive species, such as the zebra mussel, on domestic

water supplies, however, can be predicted with a high

degree of precision.

Research Needs

1. Long-term studies are needed on the impacts of both

reintroduced species and exotic species on habitat and

water quality.

2. Additional research should address the losses of other

species and reductions in water quality after introduction

of exotic or previously extirpated species.

Fish and Aquatic Organisms
173

Key Points

The current body of research will permit managers to

address some issues, but the long-term influence on water

quality of repeated measures to control exotic species is

unknown.

Physical Habitat

Issues and Risks

The literature on stream habitat improvement is large and

diverse, ranging from simple handbooks and pamphlets

designed for volunteers to more detailed treatments for

biologists and other professionals, e.g., Hunter (1991). Most

habitat improvements are designed to slow or redirect water

flow or to create pools. Habitat improvements include

engineered structures, such as k-dams, wing dams, and

deflectors. Unless they are constructed with chemically

preserved materials (such as creosote or pressure-treated

wood), the materials used in most structures should not pose

a direct threat to water quality. Indirectly, structures may

cause changes in sediment storage and routing patterns by

causing excessive scouring of channel bottoms and sides.

Turbidity may increase after installation of even properly

sited structures, but structures installed without regard for

natural channel processes can cause major disruptions in

flow patterns and trigger accelerated channel erosion.

Only recently have managers attempted to mimic the

natural structure of streams by adding native materials such

as large or coarse woody debris (CWD) to streams. For

purposes of scientific discussion, CWD includes any piece

of wood that is at least 4 inches [10 centimeters (cm)] in

diameter and 3 to 4 feet (0.9 to 1.2 m) long (Bisson and

others 1987, Dolloff 1994, Harmon and others 1986, Maser

and Sedell 1994). In practice or application, managers

typically consider CWD to be wood that is at least 12

inches (31 cm) in diameter with length equal to the width of

the receiving stream channel. Logging residue or slash is

excluded from most habitat enhancement projects because it

tends to be unstable in all but the smallest stream channels.

Woody debris enters stream channels naturally by a number

of routes, including bank undercutting, windthrow, and as a

result of catastrophic events, such as snow and debris

avalanches and hurricanes. Managers add CWD by direct

felling or toppling of streamside trees or by transport from

more distant sources.

Findings from Studies

In general, the relatively small amounts of wood added by

managers to enhance or restore stream habitat are not likely

to exert a major impact on water quality. There can be

exceptions when pieces dislodged during floods plug

culverts, bridge openings, or other structures and cause

accelerated erosion or the failure of streamside roads or road

crossings. When large amounts of fine debris from bark and

branches accumulate, dissolved oxygen may be depleted and

hydrogen sulfide and ammonia produced (Sedell and others

1991). Leachates from logs may contain toxins, but they are

unlikely to reach significant concentrations under natural

conditions (Schaumburg 1973, Thut and Schmeige 1991).

Reliability and Limitations of Findings

Since intensive research began about 20 years ago, a large

body of information has accumulated on the ecology and

management of CWD in streams. Many studies, including

historical observations and experiments, have been conducted in watersheds across North America and several

other continents. Reliability is high.

Studies have demonstrated the benefits of large wood in

streams. Whether in small streams or larger rivers, research

suggests that wood in the water is good for fish and, except

in certain well-defined situations, not detrimental to

domestic water supplies.

Research Need

Research is needed to evaluate the influence of woody

vegetation planted to stabilize fish habitat. For example, use

of black locust or other nitrogen-fixing plants may increase

nitrogen content of the water, resulting in either excessive

algal blooms or the need to remove the nitrogen prior to

domestic use.

Key Point

The information available should allow managers to address

most issues related to the effects on drinking water of

woody materials installed to improve fish habitat.

Chapter 17
174

Liming of Acidified Waters

Issues and Risks

Water from areas where the bedrock does not have a high

buffering capacity can be acidified by major soil disturbing

activities such as road building, mining (Nelson and others

1991), or acid precipitation. Associated problems include

disruption of physiological processes for many aquatic

organisms and increased concentrations of toxic forms of

metals such as aluminum. Left uncorrected, continuing

acidification can kill entire faunas (Olem 1991). The quality

of surface water can be dramatically improved by treating

streams, lakes, or whole catchments with a soluble basic

mineral such as crushed limestone.

Findings from Studies

The beneficial effects of liming have been known for

hundreds of years (Henrikson and Brodin 1995, Porcella and

others 1995). Liming, however, is at best a temporary

solution and must be repeated to maintain water quality. The

primary benefit of liming has been to preserve or recover

fish stocks and diversity of aquatic species. Liming generally increases pH and reduces concentrations of metals,

including toxic forms of aluminum (Wilander and others

1995). Concentrations of many other trace metals also

decreases as pH increases (Vesely 1992). Liming may cause

the precipitation of metals and stress to aquatic species

present in mixing zones, such as in downstream reaches

where acidic tributaries join with limed water (Henrikson

and Brodin 1995). The effects of liming are nearly always

positive; however, limestone sand or gravel applied directly

to a stream may contribute to the sediment load during high

flows.

Reliability and Limitations of Findings

In the last 20 to 30 years, much research has described the

mechanism, consequences, and treatment of acidified

surface water. The findings are considered highly reliable.

Impacts of liming on water quality have often been studied,

and the findings are widely applicable.

Research Needs

In general, knowledge of virtually all of the long-term (more

than 20 years) effects of liming on aquatic ecosystems and

water quality is lacking. Henrikson and Brodin (1995)

compiled a comprehensive list of research questions about

liming. In the area of water quality, some of the topics

needing particular attention include the effects of liming on

the uptake and transport of mercury and other metals in

aquatic organisms and the consequences of reacidification if

liming is stopped.

Key Point

For treatment of individual small watersheds, current

knowledge should permit managers to address most issues

related to sources of drinking water associated with liming.

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Fish and Aquatic Organisms
177

Part V:

Effects of Mining and

Oil and Gas Development

on Water Quality

Greens Creek Mine, Tongass National Forest, Alaska, with sediment pond.

Photo by Stephen Glasser
179

Chapter 18

Hardrock Mining

Mike Wireman1

Introduction

Mining can significantly impact the quality of water used

for domestic and municipal water supplies. These impacts

can be brief or long lasting, and they differ with the type of

ore, the mining method, the method of ore processing, the

effectiveness of water management, and after mining ceases,

the overall nature of mine closure. The impacts include

transport and deposition of sediment, acid runoff, and

release and transport of dissolved metals and other associated mine contaminants.

Hardrock mining is defined as the extraction of precious and

industrial metals and nonfuel minerals by surface and

underground mining methods (Lyon and others 1993). In the

United States, extensive hardrock mining started in the

1880’s, and, for the next 70 to 80 years, was a major

industry in many States. Many metals and minerals produced by hardrock mining are valuable natural resources

and have been important to the economy of many States.

The legacy of the active period of hardrock mining includes

more than 200,000 abandoned or inactive mines. As of

1992, there were more than 500 operating mines in the

United States, of which, more than 200 are gold mines. As

of 1997, there were approximately 60 mine sites in 26 States

on the Federal Superfund National Priorities List because of

serious pollution problems.

Hardrock mining is a large-scale activity that typically

disturbs large areas of land. Unlike other industrial facilities,

mines must be located at specific places where ore bodies

are found. Many ore bodies and mines are located on public

land administered by Federal land management agencies—

the Forest Service in the U.S. Department of Agriculture

and several agencies in the U.S. Department of the Interior.

Mines on public land are frequently located in watersheds with relatively little development. Unless proper

environmental controls are used during mining and ore

processing, and after mine closure, serious environmental

damage can result. During the first half of the 20th century,

environmental controls were very limited or nonexistent

and, as a result, numerous abandoned mines continue to

cause serious environmental damage. Ownership of abandoned mines on public land is often difficult or impossible

to establish. To date, the Forest Service does not have a

complete inventory of these mines. However, some State

mine-permitting agencies have compiled inventories.

Because of the high waste-to-product ratios associated with

mining most ore bodies, large volumes of waste are generated. Mine waste includes all of the leftover material

generated as a result of mining and ore processing activities.

Most mine waste is considered to be nonmarketable, but

mine waste materials often contain environmentally

significant concentrations of heavy metals and precious

metals.

This report describes the major potential impacts on the

quality of public drinking water sources associated with the

various elements of mining. It is recognized that some

discussion may not accurately reflect the environmental

conditions at modern hardrock mining operations that are

well designed, operated, and regulated. The intent of the

discussion is to describe environmental problems that may

occur at historic, current, and future mine sites.

Mining Methods

Precious metals and industrial metals typically occur in

disseminated ore bodies or vein deposits. The two primary

methods used to mine metals and minerals include surface

or open-pit mining and underground mining. Surface or

open-pit mining is typically used for large shallow ore

bodies, which have a low metal or mineral value per volume

of rock. Underground mining is typically used when the

mineralized rock is deep and occurs in veins.

Surface or open-pit mining often requires the removal and

disposal of soil and rock overburden that contains no target

mineral. The underlying ore body typically includes some

rock that contains uneconomical concentrations of the target

mineral. This waste rock is also removed and typically 1

 Regional Ground-Water Expert, U.S. Environmental Protection Agency,

Region VIII, Denver, CO.
180

Hardrock Mining

stockpiled or otherwise disposed of. The portion of the ore

body to be mined is drilled, blasted, and transported to a

facility where it is crushed and prepared for milling or

leaching.

Underground mining requires the excavation of vertical

shafts, horizontal adits, and inclined adits to access the ore

body. The rock that is excavated during adit construction is

commonly referred to as development rock. Once the ore

body is reached, horizontal passages called drifts and

crosscuts are developed on numerous levels and the ore is

mined. Waste rock and ore are transported to the surface via

rails or small trucks, or they are hoisted to the surface in

vertical shafts.

In both surface and underground mining, extraction of ore

waste materials requires heavy equipment and explosives. A

commonly used explosive is a mixture of ammonium nitrate

and fuel oil. As the development rock, waste rock, and ore

are removed, they are typically transferred to large trucks

for transportation to storage or processing facilities. Overburden and development rock usually do not contain

minerals that reduce the quality of surface or ground water,

and they can be used as mine backfill, but they are typically

disposed of in piles near the mine site. Waste rock from the

ore body can contain environmentally significant amounts

of metals and should be tested for acid-generating potential.

It is important to segregate waste rock that is potentially

acid generating and not use it for mine backfill or impoundment dams.

Surface and underground mines typically extend below the

local and regional water table or both. As a result, ground

water may flow into the mine pit or underground workings.

Water collecting in the mine pit or workings must be

removed. In open-pit mines, this water is typically pumped

out and discharged to nearby surface water. In underground

mines, the water can be pumped out or drainage adits can be

constructed at or below the lowest mine level to allow for

free drainage of the water entering the workings. Many

precious metal ore bodies occur in mountainous terrain

where the host rock is commonly comprised of igneous or

metamorphic rocks. In these types of rocks, ground water

occurrence and flow is controlled by the distribution and

orientation of fractures, joints, and faults. In these settings,

ground water inflow into mine workings occurs only where

the mine workings intersect water-bearing structures.

Ore Processing

Ore processing, or milling, refers to the processing of ore

rock to create the size of the desired product, remove

unwanted constituents, and concentrate or otherwise

improve the quality of the desired product. Applicable

milling processes are determined based on the physical and

chemical properties of the target metal or mineral, the ore

grade, and environmental considerations.

Amalgamation

This is the process where metallic mercury is added to gold

ore to separate the gold from the ore rock. When liquid

mercury comes in contact with gold, it bonds with the

surface of the gold particles (amalgamation). The mercurycoated gold particles coalesce or collect into a gray, plastic

mass. When this mass is heated, the mercury is driven off

and the metallic gold remains.

Flotation

The physical and chemical properties of many minerals

allow for separation and concentration by flotation. Finely

crushed ore rock is added to water containing selected

reagents. These reagents create a froth, which selectively

floats some minerals while others sink. Common reagents

include copper, zinc, chromium, cyanide, nitrate, phenolic

compounds, and, for copper ore, sulfuric acid. The waste

(tailings) and the wastewater are typically disposed of in

large, constructed impoundments.

Leaching

Leaching refers to processes that involve spraying, pouring,

or injecting an acid or cyanide solution over crushed and

uncrushed ore to dissolve metals for later extraction. The

type of solution used depends on the ore’s physical and

chemical characteristics. Leaching is used almost exclusively on low-grade ore. The main types of leaching include

dump, heap, and in situ leaching. For each type, a nearby

holding area (typically a pond) is used to store the pregnant

solution prior to recovery of the desired metal by a chemical

or electrical process. Once the desired metal is recovered,

the solution is reused in the leaching process.

In dump leaching, the material is generally piled on the

ground, and the leaching solution is applied to the pile by

spraying, injecting, or washing. Dump leach piles can be

very large, often covering hundreds of acres (hectares) and

containing millions of tons of ore rock. Leaching solutions

aided by precipitation dissolve the desired metals. Dump

leach piles are not placed on clay or synthetic liners. The
181

pregnant solution drains away from the bottom of the leach

pile to a holding pond. Pregnant solution can be lost to the

subsurface, which reduces the amount transported to the

holding pond, and potentially contaminates ground water.

Dump leaching is used for very low-grade ore.

Heap leaching is used for higher grade ores and is generally

conducted on a smaller scale than dump leaching. The ore is

usually crushed and is placed on a pad constructed of

synthetic materials or clay. These low-permeability liners

help maximize recovery of the leachate.

In situ leaching involves pumping a reagent (commonly a

sulfuric acid solution) directly into the ore body. The reagent

dissolves the desired mineral, and the pregnant solution is

collected and pumped to the surface for extraction of the

desired mineral.

Leaching can recover economic quantities of the desired

mineral for months, years, or decades. When leaching no

longer produces economical quantities of metals, the spent

ore is typically rinsed to dilute or otherwise detoxify the

reagent solution to meet environmental standards. If

standards are met, the rinsing may be discontinued and the

leached material may be allowed to drain. The spent ore is

then typically left in place.

Water Management

Management of water at large mine sites is a critical element

of a mining operation. At large mine sites that include a mill

and a tailings impoundment, water management is difficult.

It is complicated by the many management requirements,

which may include the dewatering of open pits and underground mine workings or both, the transportation of surface

runoff across mine sites, the use and containment of water

used for ore processing, and the need to meet applicable

water-quality standards for all discharges from the mine

site. Historically, the management of water has not focused

enough on prevention of environmental impacts. Nationwide, there have been numerous incidents where contaminated water from a mine site has been improperly

discharged, impairing the quality of surface water.

Waste Management

Hardrock mining typically produces large volumes of solid

waste, including overburden, development rock, waste rock,

spent ore, and tailings. Overburden, development rock, and

waste rock are typically stockpiled at the mine site. Some of

these materials may be used as pit backfill or uncommonly

for backfill of underground workings. Overburden and

development rock usually pose minimal threats to the

environment. Waste rock can contain significant concentrations of metals and pyrite and may present an environmental

problem. Some waste rock stockpiles may be left in place

for future ore processing.

Tailings are the waste solids remaining after ore processing.

Tailings generally leave the mill as slurry consisting of 40 to

70 percent liquid and 30 to 60 percent fine-grained solids.

Tailings can contain significant concentrations of heavy

metals and other contaminants. Most tailings are disposed of

in on-site impoundments. Historically, tailing impoundments were not lined and were located without consideration

of potential environmental impacts. Modern tailing impoundment design often includes low-permeability clay or

synthetic liners, engineered caps designed to eliminate or

minimize infiltration of water into the tailings, and collection systems to capture leachate that escapes from the

impoundment.

Seepage from tailing impoundments is often unavoidable

and raises the probability of surface water and ground water

contamination. Such seepage and acid rock drainage may

require water treatment long after the active life of the

facility. Failure to maintain adequate hydrostatic pressure

within and behind an impoundment dam may result in

failure of the impoundment structure, releasing tailings and

effluent to surface and ground water.

Spent ore is a waste material that is generated at mines that

utilize heap or dump leaching. The volume of spent ore can

be very large and can contain environmentally significant

residual amounts of leaching reagent and dissolved metals.

Both spent ore and tailings need to be actively managed for

years after mine closure to ensure that leachate does not

escape to a nearby stream or infiltrate into underlying

ground water.

Mine Closure

Closure of a mining operation occurs during temporary

shutdown of operations or permanent decommissioning of

the facilities. Depending on the type of mine, the size and

nature of the area of disturbance, and the type of ore

processing, active management of the mine site including

water management may be necessary for years or even

decades following closure. Until recently, reclamation was

limited to grading and revegetating waste materials and pits

to minimize erosion and improve the visual landscape.

Permanent closure now routinely includes some or all of the

following: removal and disposal of stored fuels and

Chapter 18
182

Hardrock Mining

chemicals, structure tear down, removal of unnecessary

roadways and ditches, shaft and adit plugging, waste

detoxification, capping of tailings, backfilling pits, and

active water management to ensure that all applicable waterquality standards are met. In numerous cases, a water

treatment facility must be operated and maintained. At mine

sites where acid mine drainage is a problem, water treatment

may be necessary for decades.

The long-term nature of mining impacts may require that

environmental monitoring (source, early warning, and

compliance monitoring), contingency planning, and financial insurance be in place for decades. Geochemical conditions in the ore body, waste rock, tailings, and workings can

change over time. Hence, the ability is needed to make

necessary changes in water control and water treatment after

mine closure.

Issues and Risks

At hardrock mines, adits and shafts, underground workings,

open pits, overburden, development rock and waste rock

dumps, tailings impoundments, leach pads, process ponds,

and mills are known sources of heavy metals, sulfate,

cyanide, and nitrate. If released in environmentally harmful

concentrations, these contaminants can have significant

negative effects on the quality of surface water and ground

water for public drinking water sources. Dissolved and total

metals concentrations can impact public water supplies and

the aquatic health of stream and riparian systems.

Surface runoff is a key mechanism for release of pollutants

into streams and lakes. Seepage from tailings ponds and

waste rock piles, unwanted releases from process water

ponds or wastewater ponds, drainage from underground

workings, and discharge of pit water may contaminate water

resources. Surface waters may also be impacted by contaminated ground water or contaminated by heavy metals in

sediments. The mobility of contaminants is increased by

exposure to rain and snowmelt.

A variety of complex geochemical and hydrogeological

processes control the transport, attenuation, and ultimate

distribution of heavy metals and other mine-related contaminants. Dissolved and suspended contaminants are transported to aquifers and streams via complex overland and

subsurface pathways. This complexity, combined with the

large scale of mining activities and the numerous minerelated sources of contaminants, make water-quality

assessments and restoration and remediation of mine sites

very difficult.

Environmental problems are often more difficult to deal

with at abandoned mine sites that lack environmental

monitoring. Several thousand abandoned and inactive mines

exist on public land. The U.S. Department of Agriculture,

Office of Inspector General estimates that there are more

than 38,000 abandoned and inactive hardrock mines on land

administered by the Forest Service.

The major types of water-quality impacts include erosion

and sedimentation, acid rock drainage, cyanide leaching,

and dissolution and transport of toxic metals. These impacts

are discussed in the following sections.

Erosion and Sedimentation

Because mining may disturb large areas and expose large

quantities of earthen materials, erosion and subsequent

transport of sediment to surface water can be a major

concern. Major sources of erosion and sedimentation

include open-pit areas, heap and dump leach piles, overburden, development and waste rock piles, tailings piles and

dams, haul and access roads, ore stockpiles, vehicle and

equipment maintenance areas, exploration areas, and

reclamation areas. Historically, erosion and sedimentation

have built up thick layers of mineral fines and sediment in

floodplains and streams at many mine sites. These sediments can carry attached chemical pollutants and toxic

metals, which can be stored in floodplain and bed sediments. To avoid these problems, erosion and sedimentation

must be controlled from the beginning of operations through

postclosure treatments.

Sediments and minerals deposited in floodplains can impact

the quality of nearby surface water and underlying ground

water. Oxidation of sulfide minerals may lower the pH of

surface runoff, thereby mobilizing heavy metals that can

infiltrate into underlying ground water and/or be transported

to nearby surface water. Reduced soil pH also may kill

riparian vegetation.

Drinking water impacts associated with erosion and sedimentation are discussed in chapter 2.

Mining disturbances also can increase surface runoff, which

can result in increased streamflow velocities and volumes,

downstream flooding, scouring of stream channels and

structural damage to water diversions, drinking water

intakes, bridge footings, and culverts.
183

Acid Rock Drainage

A major water-quality problem at hardrock mine sites is the

formation of acid rock drainage and the associated mobilization of toxic metals, iron, sulfate, and total dissolved solids.

The formation of acid rock drainage results from the

exposure of sulfide minerals (pyrite, pyrrhotite, galena,

sphalerite, and chalcopyrite) to air and water. Sulfide

minerals are commonly associated with coal deposits and

precious and heavy metal ore bodies. Pyrite (FeS), the most

common sulfide mineral, reacts with water and oxygen to

produce ferrous iron (Fe+2), sulfate (SO4), and acid (H+). In

waters where oxidizing conditions are prevalent and the pH

is >3.5, ferrous iron will oxidize to ferric iron. Much of the

ferric iron precipitates as iron hydroxide. Some ferric iron

remains in solution and continues to chemically accelerate

the oxidation of pyrite and subsequent generation of acid. As

the pH continues to decrease, the oxidation of ferrous iron

decreases and the precipitation of iron hydroxide decreases.

This results in a greater dissolved concentration of ferric

iron and, therefore, a greater rate of sulfide (pyrite) oxidation. The oxidation of sulfide minerals is also catalyzed by

Thiobacillus ferrooxidans bacteria. These bacteria, which

are common in the subsurface, can increase the rate of

sulfide oxidation by 5 or 6 orders of magnitude. When low

pH water comes in contact with metal-bearing rocks and

minerals, a number of toxic metals dissolve and are transported by the water. Different metals are dissolved over

different ranges of pH. The most common metals associated

with sulfide minerals include lead, zinc, copper, cadmium,

and arsenic.

Both water and oxygen are necessary to generate acid

drainage. Water is both a reactant and a medium for the

bacteria that catalyze the oxidation process. Water also

transports the oxidation reaction products and the associated

dissolved metals. Atmospheric oxygen is a very strong

oxidizing agent and is important for bacterially catalyzed

oxidation at pH values below 3.5. Surface water and shallow

ground water typically have relatively high concentrations

of dissolved oxygen.

Acid rock drainage can be discharged from underground

mine workings, open-pit walls and floors, tailings impoundments, waste rock piles, and spent ore from leaching

operations. Acid rock drainage occurs at both active and

abandoned mines. Acid generation and drainage of acid

water with high concentrations of dissolved metals affect

both surface and ground water. Ingesting water contaminated by heavy metals can have significant health affects for

humans and aquatic organisms, including water birds and

fish. Metals and other mine-related contaminants in sources

of drinking water can exceed water-quality standards.

Expensive treatment or acquisition of another source of

water may be the only alternatives.

Cyanide Leaching

For over a century, cyanide has been used as a pyrite

suppressant in base metal flotation and in gold extraction.

Dump leaching and heap leaching operations commonly use

cyanide in the leaching solution. Continued improvements

in cyanide leaching technology have allowed the economic

mining of lower grade ores. As a result, increasing amounts

of cyanide are being used in mining. The mining industry

now uses most of the sodium cyanide used in the United

States. More than 100 million pounds (45 million kilograms) were used by gold and silver leaching operations in

1990.

Cyanide can cause two major types of environmental

impacts: (1) ponds and ditches (and to a lesser degree,

tailings impoundments) that contain process water containing cyanide solutions can present an acute hazard to

wildlife, especially aquatic birds; and (2) spills or other

unwanted releases of cyanide solution from ponds, leach

impoundments, spent ore piles, or tailings impoundments

can enter surface water killing fish and contaminating

drinking water sources. During the 1980’s and early 1990’s

as the use of cyanide leaching increased worldwide, a

number of serious cyanide spills and unwanted releases have

occurred. Impacts on wildlife and streamwater quality have

been significant. These incidents and the acute toxicity of

cyanide have focused public attention on the use of cyanide

in the mining industry.

When cyanide is inhaled or ingested, it interferes with an

organism’s oxygen metabolism and can be lethal in a short

time. Cyanide is much more toxic to aquatic organisms than

to humans. The acute aquatic standard is 22 milligrams (mg)

per liter and the chronic aquatic standard is 5.2 mg per liter.

The maximum contaminant level for public drinking water

supplies is 200 mg per liter. These values are for total

cyanide even though toxicity is caused by free cyanide.

Total cyanide is usually measured because it is difficult to

measure free cyanide. Nitrate, a breakdown product of

cyanide, is also a drinking water problem (see chapter 2).

Cyanide that is dissolved in water readily complexes with

metals. At pH values below 9, weaker cyanide compounds

can dissociate and hydrogen cyanide (HCN), a volatile

poisonous gas, is formed as a byproduct. If cyanidecontaminated water infiltrates into unsaturated soil and the

pH of the water is lowered to below 9, free cyanide can

volatilize to hydrogen cyanide. Cyanide can also be attenuated to some degree by other processes, including adsorption, precipitation, oxidation to cyanate, and biodegradation.

Chapter 18
184

Hardrock Mining

Once the leaching of ore dumps or heaps is complete, it is

necessary to rinse the spent ore until the appropriate cyanide

standard is reached. In arid regions, getting enough water to

rinse heaps or dumps can be a significant problem. In wet

climates, excess water from heavy precipitation can increase

the risk of unwanted cyanide releases from leach dumps or

heaps. The chemistry of the spent ore and the associated

water in leaching impoundments can change over time,

creating a potential for continued release and transport of

dissolved metals long after the cyanide concentration has

been reduced by rinsing. Factors affecting the chemistry of a

heap leaching impoundment include pH, moisture, and ore

mineralogy.

Also of significant concern is the long-term structural

stability of large heap leach impoundments. The physical

characteristics of the leached ore, the physical configuration

of the impoundment, and specific site conditions affect the

long-term structural stability of a leach impoundment.

Structurally unstable impoundments may fail, allowing

contaminated leachate or sediments to reach public drinking

water sources.

Transport of Dissolved Contaminants

Dissolved contaminants (primarily metals, sulfate, and

nitrate) can migrate from mining operations to underlying

ground water or nearby surface water that is a source for

drinking water. Discharges of process water, mine water,

runoff, and seepage from mine waste piles or impoundments

can transport dissolved contaminants to source water.

Under specific conditions, dissolved constituents in surface

water can precipitate and attach to sediments. Elevated

concentrations of lead and mercury are often found in

sediments while being undetected in the water column.

Sediment contamination may affect human health through

consumption of fish that bioaccumulate toxic pollutants.

Contaminated sediment provides a long-term potential

source of pollutants that, under certain geochemical conditions, can dissolve in the water column.

The likelihood of contaminants dissolving and migrating

from mine waste materials or mine workings to ground

water depends on the nature and management of the waste

materials, the local hydrogeologic setting, and the geochemical conditions in the underlying vadose zone and

aquifer. Risks to human health and the environment from

contaminated ground water can be significant. In many

hydrogeologic settings, ground water discharge provides a

significant percentage of stream baseflow. In this manner,

ground water contaminated by mining activities can also

contaminate surface water.

At some locations, naturally occurring substances in an ore

body can be a significant source of contaminants. The rocks

that comprise ore bodies contain varying concentrations of

nontarget minerals, including radioactive minerals. Other

minerals may be present at concentrations that can be toxic

and can be mobilized by the same geochemical and hydrological processes that control transport of mine-related

contaminants. Nontarget minerals that can pose a risk to

drinking water sources include aluminum, arsenic, asbestos,

cadmium, chromium, copper, iron, lead, manganese,

mercury, nickel, silver, selenium, thallium, and zinc. Unlike

many other types of industrial operations and associated

discharges, contaminant loading from hardrock mine sites

can vary significantly with the season.

Findings from Studies

During the past 10 years, an increasing number of environmental studies have characterized the environmental impacts

associated with active, inactive, and abandoned hardrock

mines. Most of these studies have focused on water-quality

impacts. In 1995, the Bureau of Land Management, the U.S.

Geological Survey (USGS), and the Forest Service jointly

developed a strategy to address cleanup of abandoned mines

on Federal land (Nimick and von Guerard 1998). As part of

this strategy, the USGS developed an abandoned mine land

initiative that included numerous pilot studies in the Boulder

River watershed in Montana and the Animas River watershed in Colorado. Most of the applied research efforts

associated with this initiative were aimed at determining

sources and magnitudes of metal loadings in nearby streams.

A number of these studies documented significant metal

loading from mining-related facilities and also from

unmined areas underlain by sulfide ore bodies.

Using authorities under the Clean Water Act and Comprehensive Environmental Response, Compensation and

Liability Act (CERLA), the U.S. Environment Protection

Agency (EPA) has conducted a number of studies in

Colorado and Montana. These studies have characterized the

hydrologic pathways and geochemical processes that control

the release and transport of toxic metals from mining

facilities to underlying ground water and nearby surface

water. Allen and Stanley (1998) summarize water-quality

data collected in 1974–97 from streams that flow out of the

New World Mining District in southwestern Montana. Water

quality in two different streams has been significantly

impacted by metals loading from mine workings, mine

waste, and, to a lesser degree, by “natural” background
185

loading (metals mobilized and transported to streams in the

absence of any mine-related disturbance). In Daisy Creek,

which flows past a mine pit that has been backfilled with

mine waste, dissolved copper concentrations have ranged

from 0.93 to 6.22 mg per liter at a location just downstream

from the mine pit. The average concentration in 13 samples

was 2.24 mg per liter. The drinking water standard is 1.0 mg

per liter, and the chronic aquatic standard is 0.012 mg per

liter. Dissolved iron concentrations at the same location

ranged from 0.55 to 12.30 mg per liter. The average concentration in 13 samples was 3.53 mg per liter. The drinking

water standard for iron is 0.30 mg per liter.

Studies conducted by EPA’s Region VIII and the Colorado

Division of Minerals and Geology in the Chalk Creek

mining district in southern Colorado have documented

extensive metal loading to Chalk Creek from the historic

Mary Murphy gold mine. Zinc loading attributed to the

extensive underground workings in Chrysolite Mountain

and mine waste piles in the floodplain of Chalk Creek has

been well documented. Zinc concentrations as high as

192,300 micrograms (μg) per liter have been measured in

leachate from an old tailings pile less than one-fourth of a

mile from Chalk Creek.2 Data from 1999 indicate excessive

zinc levels at three locations: (1) as high as 32,730 μg per

liter in water discharging from the portal of the Golf Tunnel,

which is the lowermost adit in Chrysolite Mountain;

(2) 221,300 μg per liter in ground water seeping down

through the upper workings in Chrysolite Mountain; and

(3) 341 μg per liter in Chalk Creek below the Mary Murphy

mine.3 The drinking water standard for zinc is 5,000 μg per

liter and the chronic aquatic standard is 110 μg per liter (at

100 mg-per-liter hardness). It is clear from these data that

mining activities have had a significant impact to ground

water and surface water in the vicinity of the Mary Murphy

mine.

Reliability and Limitations of Findings

Data and information on potential environmental impacts

related to hardrock mining have increased greatly in the past

10 years. Numerous investigations and published reports

have documented movement of toxic metals to ground water

and surface water from mines and mine-related facilities.

The data from the increasing number of reports is reliable

because the findings are comparable and often present the

same conclusions. Many of the study results have been

published in peer-reviewed literature.

One point of disagreement and uncertainty is the significance of “natural” background metal loadings versus metal

loadings that result from mining activities. A number of

studies have attempted to separate “natural” from mancaused loading (Nimick and von Guerard, 1998). Researchers have used water-quality data, including isotopes and

tracers, to try to identify loading caused by leaching of

unmined ore bodies. However, to date there has been no

reliable technique developed to clearly separate natural from

man-caused loading.

Research Needs

1. Research needs related to the environmental management of hardrock mine sites include two primary areas:

(a) characterization of hydrologic and geochemical

processes that control the release and transport of minerelated contaminants away from a mine site to ground

water or nearby surface water; and (b) development of

workable, passive systems for treating water with low pH

and high concentrations of dissolved metals.

2. Hardrock mines often occur in complex hydrogeologic

settings where a standard approach to characterization of

ground water and surface water is inadequate. A mine can

greatly disturb natural hydrologic systems, creating major

water pollution problems. It is critical that we continue to

improve characterization approaches and tools. An increased understanding of processes, which control

distribution of mine-related contaminants, will be helpful

for planning future mines and implementing effective

environmental controls.

3. Capital, operating, and maintenance costs associated with

active treatment of contaminated mine water are prohibitive at most mine sites. It is extremely important to

continue research directed at developing efficient and

cost-effective passive treatment technologies that can be

operated year-round at high elevations. Research must

continue on the use of organic substrata to facilitate the

utilization of sulfide-reducing bacteria to remove dissolved metals from water. To date these technologies

have been limited by the inability to deal with high-flow

rates and the extreme climatic conditions at high

elevations.

2 Science Applications International Corporation. 1993. Chalk Creek

nonpoint source project case history. 99 p. Unpublished report prepared for

U.S. Environmental Protection Agency, Region VIII, Denver, CO. On file

with: Science Applications International Corporation, 999 18th Street,

Denver, CO 80202–2405.

3

 Wireman, Mike. 1999. Unpublished field data from Mary Murphy mine—

Chalk Creek Mine District, Chaffee County, CO. [Not paged]. On file with:

Mike Wireman, U.S. Environmental Protection Agency, Region VIII, 999

18th Street, Suite 500, Denver, CO 80202–2405.

Chapter 18
186

Key Points

1. Management practices are commonly used to control

erosion and sedimentation at mine sites. The selection of

erosion control measures is based on site-specific

considerations, such as facility size, climate, geographic

location, geology, hydrology, and the environmental

setting of each mine site. Mining facilities are often in

remote locations and may operate only seasonally or

intermittently, but they need year-round pollution

controls. At least six categories of management practices

are available to limit erosion and the off-site transport of

sediment including discharge diversions, drainage and

stormwater conveyance systems, runoff dispersion,

sediment control and collection, vegetation and soil

stabilization, and capping sources of contamination.

2. No easy or inexpensive solutions to acid rock drainage

are currently available. An appropriate approach is to

isolate or otherwise segregate waste with acid-generation

potential, and then treat them appropriately. Management

may include minimizing contact with oxygen and water

and/or neutralizing acid that is produced with natural or

introduced material. Techniques used include subaqueous

disposal, covers, waste blending, hydrologic controls,

bacterial control, and treatment.

3. Acid-generation prediction tests are increasingly relied

upon to assess the long-term potential of pit walls and

floors, underground workings, and mine waste to

generate acid. Mineralogy and other factors affecting the

potential for acid rock drainage are highly variable from

site to site, and this can result in less than accurate

predictions. In general, the methods used to predict the

acid-generation potential are classified as either static or

kinetic. Static tests are intended only to predict the

potential to produce acid rather than predict the rate of

acid generation. Static tests can be conducted quickly and

are inexpensive compared with kinetic tests. Kinetic tests

are intended to mimic the processes found in the environment of the ore body or waste unit environment; however, they require more time and are more expensive than

static tests. Reliable dynamic tests that are faster and less

expensive are needed.

4. The heightened awareness of the potential environmental

problems associated with cyanide leaching led Federal

land managers and States to implement increasingly

stringent regulations and guidelines. These regulations

and guidelines address the design of facilities that use

cyanide and include requiring or recommending use of

liners with heap leach piles or tailings impoundments,

monitoring of solutions in process waters and ponds,

treatment requirements for cyanide-containing wastes,

and closure and reclamation requirements. Operators are

generally required to take steps either to reduce or

eliminate unwanted releases of cyanide solutions or to

reduce cyanide concentrations in exposed materials to

below standards. Regulatory requirements and guidelines

on the allowable concentration of cyanide in exposed

process solutions vary. When numeric limitations are

established, they generally range around 50 mg per liter.

Literature Cited

Allen, Bryan K.; Stanley, Dan R. 1998. Site assessment and sampling

activities report—New World Mine, Cooke City, Montana. URS

Operating Services. Contract 68–W5–0031. Washington, DC: U.S.

Environmental Protection Agency. 108 p.

Lyon, J.S.; Hilliard, T.J.; Bethell, T.N. 1993. Burden of gilt. Washington,

DC: Mineral Policy Center. 24 p.

Nimick, David A.; von Guerard, Paul, eds. 1998. Science for watershed

decisions on abandoned mine lands; 1998 February 4–5; Denver. U.S.

Geo. Surv. Open File Rep. 98–297. Washington, DC: U.S. Government

Printing Office. 71 p.

Hardrock Mining
187

Chapter 19

Coal Mining

Mike Wireman1

Introduction

The mining of coal can have many of the same environmental impacts to water quality as hardrock mining. However,

some aspects of coal mining are different enough to warrant

a separate discussion. After a brief description of coal

mining, this chapter focuses on aspects of coal mining that

are significantly different from hardrock mining with regard

to the potential to impact water quality.

Coal accounts for one-third of the total energy usage and

more than one-half of the electricity generated in the

country (U.S. Geological Survey 1996). Domestic coal

production has been steadily increasing since the 1950’s. In

1998, total domestic production was 1.18 billion tons.2

Approximately 570.5 million tons were produced in States

east of the Mississippi River and 547.6 million tons from

States west of the Mississippi River. Coal production in the

West has almost doubled since the passage of the 1991

Amendments to the Clean Air Act. Wyoming leads the

Nation in coal production. West Virginia and Kentucky are

second and third, respectively. About 60 percent of domestic

production is from surface mines and 40 percent from

underground mines.

Mining Methods

Strip mining is the most common method of producing coal

from surface mines. Strip mining commonly includes the

removal and storage of topsoil, the removal of any overburden material, and the subsequent excavation of the coal

seam. As the operation advances across the land surface,

only a relatively small area is actively mined. With this

method, the overburden is removed from the advanced side

of the active mine face and placed on the retreat side, where

the coal has been mined out. There are two common

methods of underground mining: room and pillar mining

and longwall mining. In the room and pillar method, entries

or adits are driven into the coal seam, and crosscuts are

driven at right angles to the adits at spacings dictated by the

individual mine plan. A checkerboard pattern of interconnected tunnels or rooms and pillars is created. In longwall

mining, numerous crosscuts are developed around a large

block of coal. Once the crosscuts are fully developed, the

large block is completely excavated, and the chamber is

allowed to collapse. Longwall mining results in predictable

subsidence of the overlying ground surface.

Coal Preparation

Coal that is excavated from a seam or deposit requires

preparation to improve the quality and make it suitable for a

given use. Preparation includes the separation of the heavier

waste material from the lighter coal by flotation processes

that rely on the differential densities of the coal and the

waste material. Reagents are sometimes used to make the

coal more amenable to flotation. Coal preparation creates a

relatively uniform product size, reduces the amount of ash in

the coal, and may reduce the sulfur content. In addition to

clean coal, the preparation process produces a coarse,

dewatered waste rock material and a fine-grained slurry with

significant water content.

Waste Management

Waste materials are generated from coal mining and coal

preparation. Overburden material removed for surface

mining is often used to backfill the excavated area. Waste

materials from underground mining are disposed of in

mined-out workings to the extent possible, but they often

are placed in a designated waste rock disposal area on the

surface.

Large volumes of waste material can be generated from coal

preparation. Both the coarse waste rock and the fine-grained

slurry are typically disposed of in disturbed portions of the

permit area. The fine slurry waste is commonly disposed of

in an impoundment where the slurry solid settles, and the

water is reclaimed from pond on top of the impoundment.

1

 Regional Ground-Water Expert, U.S. Environmental Protection Agency,

Region VIII, Denver, CO.

2

 Personal communication. 1999. Stuart Sanderson, President, Colorado

Mining Association, 216 16th Street, Suite 1250, Denver, CO 80202.
188

Coal Mining

Environmental Regulation

With the passage of the Surface Mining and Control

Reclamation Act (SMRCA) in 1977, the coal mining

industry became the only mining sector in the United States

that is subject to mine-specific environmental regulation.

The SMCRA pertains only to coal and was promulgated by

the U.S. Congress to provide environmental standards for

reclaiming land that has been impacted by coal mining and

processing operations. The Office of Surface Mining

Reclamation and Enforcement (OSM) was established to

administer the law and regulations established by SMCRA.

The OSM can delegate the regulatory program to the State

level and most States that have substantial coal resources

have developed their own regulatory programs. Most States

have developed a permit program that regulates exploration

activities, surface mining, underground mining, and special

mining activities.

Issues and Risks

Just as in precious metal mining, the mining of coal can

result in the exposure of sulfide minerals to oxygen, water,

and bacteria. Pyrite and less commonly marcasite (FeS2) and

greigite (Fe3S4) are the primary sulfide minerals found in

coal. Oxidation of these minerals can result in the generation

of acidic water and the subsequent mobilization and

transport of heavy metals to ground water and surface water.

Mine waste and coal preparation waste can contain significant amounts of pyrite and heavy metals including cadmium, chromium, mercury, nickel, lead, and zinc. These

metals and sulfur can be concentrated in waste materials by

factors of 3 to 10 compared to raw coal (National Research

Council Committee on Accessory Elements 1979). Therefore, just as in hardrock mining acid drainage, the associated

mobilization of heavy metals in the waste materials is a

potentially significant threat to surface and ground water

resources. See chapter 18 for further discussion of acid

drainage and heavy metal mobilization.

Findings from Studies

The scientific literature includes thousands of studies on

water-quality impacts from the mining and processing of

coal. Coal mining has been much more extensively studied

in the United States than hardrock mining. In the Southeastern United States where coal mining has occurred for more

than 100 years, there are numerous documented cases of

contamination of streams from coal mining. Hyman and

Watzlaf (1997) used water-quality data from 128 different

samples of untreated coal mine drainage from mines in

Pennsylvania, West Virginia, Ohio, Tennessee, Maryland,

Montana, Kentucky, Colorado, Oklahoma, and Missouri to

characterize the occurrence of various metals and other

contaminants. Results from this study indicate that the mean

concentrations for arsenic, beryllium, cadmium, and lead

exceeded the maximum contaminant level for drinking

water and the maximum concentrations of these metals plus

antimony, chromium, and zinc exceeded the maximum

concentration level. This study also concluded that the

traditional use of manganese concentrations as an indicator

parameter for treatment thresholds is not reliable and that

water-quality protection is better achieved if individual

metal concentrations are more thoroughly considered.

Reliability and Limitations of Findings

It is clear that coal mining can mobilize and transport toxic

metals from mines and mine-related facilities to ground

water and surface water.

Research Needs

1. Within the coal mining industry, a key focus of recent

environmental research has been the environmental

effects of surface mining and power generation. A

significant amount of research has involved mining and

reclamation because these activities have the greatest

impact on the environment. Major environmental

concerns faced by the coal industry include the impacts

of surface mining on water resources and whether mined

land can be returned to productive use for crops, livestock, timber, and wildlife (White and others 1997).

Important areas of research include topsoil substitution,

reforestation, forage and row crops production, and

wetlands. All of these areas of research are aimed at

providing a better understanding of how areas that have

been disturbed by coal mining can be reclaimed to reduce

impacts on water quality.

2. An area of research that needs to be expanded is the

development of methods for characterizing the hydrologic and geochemical processes that control release and

transport of mine-related contaminants away from a mine

site to ground water or nearby surface water. This

research need is similar to that for hardrock mining. More

emphasis needs to be given to preventing or controlling

the transport of contaminants to streams.
189

Key Points

1. Mining and processing of coal clearly have the potential

to contaminate ground water and nearby surface water.

The mobilization and transport of toxic metals and other

contaminants has been well documented in many areas of

the country, especially in the leading coal-producing

States in the Southeast. In Kentucky, West Virginia,

Tennessee, and Virginia, the potential to impact surface

water quality is increased by steep topography and

narrow valleys. In this terrain, it is very difficult to mine

and process coal without impacting surface water.

2. In the Western United States, coal production has

increased significantly since 1991. In general, the western

coal has a low sulfur content, which reduces the potential

for acid rock drainage. In addition, the geologic and

topographic settings of coal deposits in Western States is

generally more amenable to the implementation of

environmental controls.

3. The SMCRA requires all coal operations to develop

environmental information, file operation and reclamation plans, and post an adequate surety prior to the

development of any coal mining operation. Management

practices are commonly used to control erosion and

sediment at mine sites. Traditionally, the focus of the

reclamation has been to restore the land disturbed by coal

mining to beneficial use. Since the passage of the Clean

Water Act, coal mining operations have been subject to

point-source permitting. However, as with hardrock

mining, no easy or inexpensive solutions to controlling

acid rock drainage are currently available. Isolating

materials will help to prevent or minimize oxygen contact

with the material and prevent water from contacting the

material.

Literature Cited

Hyman, David M.; Watzlaf, George R. 1997. Metals and other components of coal mine drainage as related to aquatic life standards. In:

Brandt, J.E., ed. Proceedings of the American Society for Surface

Mining and Reclamation 14th annual national meeting; 1997 May;

Austin, TX. Austin, TX: American Society for Surface Mining and

Reclamation: 531–545.

National Research Council Committee on Accessory Elements. 1979.

Redistribution of accessory elements in mining and mineral processing.

1: Coal and oil shale. Washington, DC: National Academy of Science.

180 p.

U.S. Geological Survey. 1996. U.S. Geological Survey programs in

Montana. USGS Fact Sheet FS–026–96. Washington, DC: U.S.

Government Printing Office. 4 p.

White, R.L.; Smith, S.F.; Miller, P.J. 1997. TU system environmental

research program. In: Brandt, J. E., ed. Proceedings of the American

Society for Surface Mining and Reclamation 14th annual national

meeting; 1997 May; Austin, TX. Austin, TX: American Society for

Surface Mining and Reclamation: 680–684.

Chapter 19
190

Chapter 20

 Oil and Gas Development

R.J. Gauthier-Warinner1

Introduction

Oil and gas exploration generally has short-term effects on

the quality of drinking water sources. Exploration consists

of geologic mapping and ground geophysical methods

consisting of surface gravity, magnetic, and seismic surveys

of the prospective area. Gravity and magnetic data are

obtained with little impact to the surface. Seismic surveys

entail the stringing of numerous arrays of geophones and the

drilling of relatively few shot holes for creating the seismic

signals. Today, the seismic energy often is generated by

thumpers mounted on large trucks, utilizing less-environmentally damaging vibroseis technology. Both methods

require a system of crude roads for access; however,

vibroseis does not require the logistical support or involve

the site disturbance that is necessary for drilling.

The Bureau of Land Management (BLM) oversees drilling

operations and specifies conditions that must be met during

drilling on public land. These conditions are designed to

meet the intent of specific laws, such as the Safe Water

Drinking Act of 1996, as well as to mitigate negative effects

on resources that may not be specifically protected under

statute or regulation.

Exploratory well drilling entails both site occupancy and

reconfiguration. It has relatively short-term effects. Exploratory wells can acquire drill cuttings and cores for visual

analysis as they probe the formation for direct information

about such rock characteristics as lithology, porosity,

permeability, and identification of pore fluids.

The majority of well drilling in today’s petroleum industry

is accomplished with rotary drills. This type of drilling

requires the circulation of a fluid to lubricate and cool the

bit, prevent plugging of the hole, and maintain the necessary

hydrostatic pressure to prevent collapse of the well. It also

counterbalances any high-pressure oil, gas, or water

encountered in any of the drilled formations. Thus, fluid

circulation helps prevent a catastrophic surge of highly

pressurized fluid, called a blowout. Blowouts can cause

fires, loss of life and property, and potential contamination

of surface drinking water sources.

The fluid circulation system uses drilling muds. Generally,

they are a water-based mixture of clays like bentonite and

inert weighting constituents like barite with special additives

mixed in low concentrations. Formulation of a particular

drilling mud is based upon downhole conditions such as

drilling depth, temperature, pressure, and the sensitivity of

an oil or gas reservoir to water. Weighting constituents are

added to the mud to counterbalance the formation pressure

and prevent the formation fluids from entering the wellbore.

The drilling mud is circulated downward through the drill

stem, into the bit, and back up the annular space between the

drill stem and the hole. It is then screened, filtered, and

recirculated through tanks back into the hole. The Forest

Service has some discretion in requiring that certain

conditions be met in fluid system design and location. There

is a risk of contamination to an intervening freshwater

aquifer. The magnitude of risk depends, among other things,

on the competence of the oil- and gas-containing rock, the

proximity of the aquifer, and the thickness and competence

of the units separating them.

Lined earthen pits, unlined earthen pits, or closed circulation

systems are used for containment of water, waste fluids

from drilling, rock cuttings, rigwash, and stormwater runoff.

Containment design is influenced by such factors as soil

conditions, depth to freshwater aquifers, proximity to

surface water sources and drainages, types of drilling fluid,

and availability of water for drilling. The design, location,

closure, and reclamation of containment systems for drilling

operations fall under the jurisdiction of both the Forest

Service and the BLM.

Upon reaching the desired depth, the well is analyzed by

electric and nuclear logs to determine whether the hole is a

potential producer. If the well is determined to have no

potential for production, it is plugged. Plugging operations

fall under the jurisdiction of the BLM and the State or both.

The responsible agency must ensure that plugging meets

local criteria for protection of underground water sources. If

the well is determined to have potential for production, the

production casing is cemented into the wellbore and the 1

 Geology Program Manager, USDA Forest Service, Minerals and Geology

Staff, Washington, DC.
191

drilling rig is replaced with a smaller completion rig. Casing

a wellbore serves several purposes. It prevents the formation

from caving into the wellbore; it provides a permanent

passageway for conveying the oil and gas to the surface; it

prevents exotic fluids from mixing with the producing

formation; and it isolates the producing zone or other

contaminating zones in the well from contact with any

freshwater aquifers penetrated by the well. Casing operations fall under jurisdiction of the BLM and the State or

both. They must be in compliance with specifications

designed to protect underground water sources and to

contain high pressures and any fluids or gases or both that

might escape to the surface and pose hazards to surface

resources, including drinking water sources.

Once drilled, cased, and completed, many wells have

insufficient force to flow without further assistance because

of material introduced by drilling or of material within the

formation itself. Two of the most common techniques of

well stimulation are acidizing and fracture treating.

Acidizing is the pumping of acid into the well to help

dissolve the impediment. When the permeability of a

reservoir is so low that it is difficult for the oil and gas to

flow into the well, the rock may be fractured to allow oil

and gas to flow freely to the wellbore. A high-pressure

fracture fluid comprised of thickened or gelled water is

pumped at high rates into the well to fracture the formation.

After a well is completed for production, the drill is removed from the site and replaced by the well head. This

phase of the operation has long-term effects because the

facilities associated with it are in place over the operating

life of the well. The Forest Service takes on long-term

responsibilities for administering ongoing operations and

monitoring conditions under which the operations occur.

Equipment design and layout are tailored to the particular

characteristics of the site and the type of production (oil,

gas, oil/gas mixtures; associated water production; oil/gas

components such as hydrogen sulfide; etc.). The emphasis is

on containment of fluids and gases, particularly in emergency circumstances. Although specific types of equipment

are continuously being designed or upgraded to provide for

environmentally safe production operations, it is often not

practical, economical, or necessary to retrofit existing

operations with some of the newer technology. The Forest

Service must work closely with the BLM and the State or

both in developing conditions of approval under which

production facilities can be safely constructed and operated.

A flowing well is any well that has sufficient pressure

belowground to cause the oil or gas to flow unassisted

through the wellbore to the surface. Artificial lift is a

technique that employs a mechanical or artificial means to

pump or lift the oil to the surface. Depending upon the

particular circumstances associated with the well, one of

several types of artificial lift can be used. Primary recovery

is the initial production of fluids using only natural sources

of energy available within the reservoir. Depending upon the

natural reservoir energy available, primary recovery can

range from <5 percent to 75 percent of the resource.

Secondary and tertiary recovery includes utilization of such

methods as injection of water, steam, carbon dioxide,

polymers, or micellar fluids to supplement natural reservoir

energy and increase fluid recovery.

Generally, oil produced from the well is a mixture of oil,

water, gas, and sand or other solid material. The sand and

other solid materials are generally removed by gravity

methods. Typically, the oil and water occur as an emulsion

and must be treated to break the emulsion. Several methods

are used for this purpose. Heaters can be used to heat the

emulsion and separate it into its oil and water constituents.

The addition of certain types of chemicals or the use of

direct current can facilitate this process.

Once at the surface, the product is transferred by gathering

lines to be treated then stored in underground or surface

tanks until it is shipped to the purchaser. Storage facilities

are comprised of welded or bolted steel tanks of various

sizes ranging from 50 barrels to more than 10,000 barrels,

depending on the scale of production. Facilities typically

include provisions for transfer to trucks or pipelines. Refer

to the discussion of roads and utility corridors in chapter 9.

Gas reservoirs generally do not contain oil, but produce gas

with varying amounts of condensate or water. They generally produce well without the addition of supplementary

energy and primary recovery methods are usually sufficient.

Recovery is often >80 percent of the resource.

Issues and Risks

The Forest Service has a limited role in administering oil

and gas operations. It has surface responsibilities only;

whereas, the BLM, the U.S. Environmental Protection

Agency (EPA), and the State have jurisdiction over subsurface operations. Additionally, the Forest Service can only

make recommendations to the BLM regarding whether or

not to issue a lease and what stipulations to apply if leased.

The BLM has no obligation to implement Forest Service

recommendations. The Forest Service must work closely

with the BLM and the State in developing conditions of

approval under which production facilities can be constructed and operations can be maintained.

Chapter 20
192

Oil and Gas Development

The Resource Conservation and Recovery Act of 1976

(RCRA), codified at 42 U.S.C. sec. 6901 et seq., conditionally exempted from regulation as hazardous wastes drilling

fluids, produced waters and other wastes associated with the

exploration, development, or production of crude oil or

natural gas. According to the EPA, exempted wastes include

well completion, treatment, and stimulation fluids; workover

wastes; packing fluids; and constituents removed from

produced water before it is injected or otherwise disposed

of. While these wastes are not considered hazardous, they

may have an effect on the quality of drinking water sources

if contamination occurs. Contamination is most likely to

occur at the surface in the event of a spill or a breach of, or

infiltration from, a containment structure.

Access roads and well pads erode and become sources of

sediment during the exploration and production phases. See

chapter 9 for discussion of roads and sediment.

Under the Clean Water Act, discharges to surface water by

oil and gas exploration and production operations are

addressed by the National Pollutant Discharge Elimination

System. Onshore discharges are prohibited except from

wells producing not more than 10 barrels per day and

discharges of produced water that are determined to be

beneficial to agriculture or wildlife (U.S. EPA 1992).

The Safe Drinking Water Act specifically addresses oil and

gas operations under its underground injection control

program. The objective of the program is to protect goodquality ground water from contamination by injected fluids.

It established a special class (class II) of injection wells for

oilfield-related fluids, the regulation of which should not

impede oil and gas production unless necessary to prevent

contamination of underground sources of drinking water. An

underground source is an aquifer that supplies drinking

water for human consumption or for any public water

system, or contains fewer than 10,000 milligrams per liter of

total dissolved solids, does not contain minerals or hydrocarbons that are commercially producible, and is situated at a

depth or location, which makes the recovery of water for

drinking purposes economically or technologically practical.

Class II regulatory programs are either directly administered

by the States under primacy programs or by EPA where

States do not administer the programs.

Injection wells are sometimes used to dispose of produced

water, a byproduct of oil and gas recovery. Most produced

water is strongly saline, with total dissolved solids ranging

from several hundred to over 150,000 parts per million

(ppm). Produced water pumped into injection wells is used

to enhance production by providing the energy needed to

drive the oil toward the producing well. Secondary recovery

may necessitate the drilling of a few to hundreds of injection

wells throughout the field, depending upon the size of the

reservoir. This water is intended to provide the energy

needed to drive the oil toward the producing well.

Secondary recovery may necessitate the drilling of a few to

hundreds of injection wells throughout the field, depending

upon the size of the reservoir.

Because produced water is beneficially recycled and is an

integral part of some crude oil and natural gas production

processes and because injection of produced water for

enhanced recovery is regulated under the Safe Drinking

Water Act’s Underground Injection Control Program, EPA

has determined that it is not a waste for purposes of RCRA

subtitle C or subtitle D.

Despite prevention measures, contamination of a drinking

water aquifer can occur as a result of improper plugging of

abandoned wells or casings, and through direct injection

into aquifers. During exploratory and development drilling,

the well has the potential to act as a conduit between

formations hosting usable aquifers and formations containing hydrocarbons, heavy metals, or chlorides associated

with accompanying brines. If the well penetrates an aquifer

and is not cased, or the casing and grouting fail, there is a

possibility for contaminants to migrate through the conduit

and into the drinking water aquifer.

Stimulation of an oil reservoir utilizing the pumping of a

fracture fluid under high pressure into the formation can

have adverse effects. If the induced fracturing extends

beyond the boundaries of the reservoir, there is a risk of

contamination to a nearby freshwater aquifer. The magnitude of risk is dependent, among other things, on the

competence of the reservoir rock, proximity of the aquifer,

and the thickness and competence of the units separating

them.

Produced water is usually a highly saline brine accompanied

by trace contaminants inherent in the reservoir. Injection of

produced water back into the reservoir for disposal or to

enhance recovery has the potential to contaminate freshwater through grout or casing failures between the injection

well and the aquifer. Injecting produced water into old

injection wells with leaking casings can introduce brine into

surface geologic strata where it can percolate to and

contaminate surface waters. Sometimes brine water is

trucked to injection wells; however, some truckers have

been known to dump the brine illegally into surface water at

stream crossings.

Corrosion or failure of any one of the numerous surface

facilities may result in leakage and subsequent migration of
193

hydrocarbons into shallow freshwater aquifers. Surface

pipes from wells to storage tanks can corrode or break and

discharge oil and brine onto the soil surface, where the

discharge can run off to streams. Pipes crossing streams can

rupture and discharge directly into streams. The degree of

contamination depends upon, among other things, the extent

and duration of leakage.

Some waste management practices associated with hydrocarbon production may have an effect on ground water. The

failure of waste pits or drilling mud pits or the utilization of

unlined pits for these purposes can allow percolation of

contaminants through the soil and into shallow aquifers.

Some natural gas contains hydrogen sulfide, carbon dioxide,

or other impurities that must be removed prior to sale.

Sweetening is the stripping of these impurities by various

chemical processes including utilization of amine, sulfinol,

iron sponge, and caustic solutions. Associated wastes may

include spent amine, gycol and sulfinol, slurries of sulfur

and sodium salts, iron sulfide and wood shavings, and

caustic filter material, which may be commingled with

produced water. These wastes may fall into a hazardous

waste category but are exempted from regulation under

RCRA. Any waste products associated with oil and gas

production, whether exempted or not, can be a risk to

drinking water sources if not managed appropriately.

The disposal of excess drilling fluid and produced water by

evaporation, road spreading, and application to the land may

have an effect on the quality of surface water. Runoff may

allow the migration of chlorides, oily wastes, or other

contaminants into streams or ground water and, thus, affect

the quality of drinking water.

Findings from Studies

With respect to the disposal by landspreading of liquid and

solid wastes, two primary concerns are their salt content and

hydrocarbon content. Studies by Deuel (1990) and Macyk

and others (1990) have shown that soil and water mixtures

or both with soluble salt levels below roughly 3,000 ppm of

total dissolved solids, exchangeable sodium percentage of

<16, and a sodium adsorption ratio of <12 cause no harm to

soil, vegetation, surface water, or ground water. Landspreading or wastes resulting in oil and grease concentrations of up to 1 percent by weight in the waste and soil

mixture or both are not harmful and will biodegrade readily.

Repetitive disking and nutrient addition can reduce concentrations in a soil mixture to these levels.

Instream monitoring by the Daniel Boone National Forest in

Kentucky revealed high concentrations of brine below oil

production well fields. In Texas, heavy sediment deposits in

streams were traced to gas well pads and service roads.

Reliability and Limitation of Findings

Anecdotal evidence of contamination or degradation of

drinking water sources from oil or gas wells exists throughout the Forest Service, particularly in areas of split mineral

estates in which the Federal Government holds surface

rights, but mineral rights are privately owned. Such estates

are most common in the national grasslands and eastern

national forests. Contamination or degradation has not been

assessed on a nationwide scale, but the level of risk depends

on the degree of monitoring and inspection. Databases

managed by the BLM and individual States may provide

more information about the extent of existing contamination

or degradation and potential for such to occur in the future.

Research Need

A quantified assessment of contamination or degradation of

surface and ground water by oil and gas operations that

draws on BLM and State data bases is needed on a nationwide scale. It would provide a more accurate framework in

which to manage oil and gas exploration and production

activities.

Key Points

All facets of oil and gas exploration and production can

affect the quality of drinking water. The Forest Service can

control the effects associated with those activities that occur

on the land surface such as site preparation, berm and pit

construction, design and location of ancillary systems, road

construction, and reclamation activities that probably have a

greater potential to affect surface water quality. The Forest

Service must work closely with the BLM, EPA, and the

States to assure that drilling, production, and waste disposal

activities are conducted so as to minimize adverse effects on

both ground water and surface water quality for public

drinking water sources.

Chapter 20
194

Literature Cited

American Petroleum Institute. 1993. Environmental guidance document:

well abandonment and inactive well practices for U.S. exploration and

production operations. 1st ed. Washington, DC: American Petroleum

Institute. 52 p.

American Petroleum Institute. 1996. Introduction to oil and gas

production. 5th ed. Washington, DC: American Petroleum Institute.

120 p.

American Petroleum Institute. 1997a. Environmental guidance

document: onshore solid waste management in exploration and

production operations. 2d ed. Washington, DC: American Petroleum

Institute. 84 p.

American Petroleum Institute. 1997b. Exploration and production:

protecting the environment. Washington, DC: American Petroleum

Institute. 24 p.

Deuel, L.E. 1990. Evaluation of limiting constituents suggested for land

disposal of exploration and production wastes. In: Proceedings of the

1st international symposium on oil and gas exploration waste management practices; 1990 September 10–13; New Orleans. Washington, DC:

American Petroleum Institute: 411–430.

Macyk, T.M.; Nikiforuk, F.I.; Weiss, D.K. 1990. Drilling waste

landspreading field trial in the Cold Lake Heavy Oil Region, Alberta,

Canada. In: Proceedings of the 1st international symposium on oil and

gas exploration waste management practices; 1990 September 10–13;

New Orleans. Washington, DC: American Petroleum Institute: 267–279.

Interstate Oil and Gas Compact Commission. 1997. Groundwater risk

assessment and remediation. Oklahoma City: Interstate Oil and Gas

Compact Commission. [Number of pages unknown].

U.S. EPA. 1992. Background for NEPA reviewers: crude oil and natural

gas exploration, development, and production. EPA Contract 68–WO–

0025. Falls Church, VA: Science Applications International Corporation,

Environmental Health Sciences Group. 85 p.

Oil and Gas Development
195

Part VI:

Implications for Source

Water Assessments and

for Land Management

and Policy

A U.S. Fish and Wildlife Service class on “working at the watershed level”

at the Potomac River near Sheperdstown, WV. Photo by Stephen Glasser
197

Chapter 21

Future Trends and Research Needs in Managing Forests

and Grasslands as Drinking Water Sources

F.N. Scatena1

Introduction

The management of forest and grassland watersheds for

drinking water supplies has been, and will continue to be, a

major activity of the Forest Service and other natural

resource agencies. However, these watersheds will continue

to support other uses, including providing timber products,

recreation, mining, fisheries, grazing, and the conservation

of biodiversity. In addition, relatively new uses like using

forests for carbon and nutrient sequestration (DeLucia and

others 1999) or the recycling of wastewater (Cole and others

1986, Sopper and Kardos 1973) will increase. The future is

also expected to bring increased competition for existing

water resources (Postel 1998) and changes from point

source to watershed-based pollution management (U.S. EPA

1997). How these watersheds will be managed in this

increasingly competitive, watershed-based, multiuse

environment will be affected by site-specific knowledge of

environmental change, technological change, and social and

administrative considerations.

Environmental Change

It is widely believed that the Earth is undergoing a period of

rapid global climate change that will significantly alter

environmental conditions in many areas during the 21st

century (Schlesinger 1997). Most global-scale climate

models predict that in the next 50 years the Earth will be

warmer, more humid, and have greater evaporation, precipitation, and runoff (Loaiciga and others 1995). However, not

all areas will be affected equally, and large areas of the

United States may actually experience more arid conditions.

Projections based on historic hydrologic conditions and

projected demands suggest that areas east of the Great

Plains and in the Pacific Northwest will have water surpluses until 2040 (Guldin 1989). In contrast, much of the

Colorado and the Rio Grande River basins, the Great Basin,

parts of California, and the lower Mississippi Valley

currently have or will have water shortages by 2040.

Moreover, 11 out of 18 water resource regions of the

Continental United States are currently diverting more than

20 percent of their streamflow for off-site uses and will be

affected by changes in either streamflow or water demand

(U.S. Department of Agriculture, Forest Service 2000). The

availability of water is predicted to seriously constrain

global food production by 2025 (Postel 1998). Water

shortages and aquifer depletion already affect many of the

World’s most important food-producing regions, including

the Western United States, northern China, the Punjab of

India, and parts of Southeast Asia, Africa, and the Middle

East.

In addition to global-scale change, local and regional

environmental change can be expected to influence municipal water supplies. These changes can occur over years or

decades and may include changes in land uses or increases

in air pollution and atmospheric inputs. Increases in upstream water withdrawals or changes in the types of water

uses (irrigation, snow making, etc.) can also modify water

quality and temporal patterns of streamflow. Changes in the

successional status of forest cover or the presence of exotic

or noxious weeds can also affect source water quantity and

quality (see chapter 11). Shifts in management practices can

also influence local environmental conditions that affect

municipal water supplies. Increased harvesting on steeper

slopes, growing wood in high-input, short-rotation plantations, and intensively managing fisheries are some examples

(see chapters 10, 17).

In response to the influence of local and regional environmental changes, assessments are now being developed to

evaluate regional risks of specific environmental hazards

(Graham and others 1991). A generic problem encountered

when doing these assessments is the lack of ecosystemspecific information on the spatial and temporal variability

of ecological and pollution-generating processes. Fortunately, technology advances in data acquisition and in the

management of spatially explicit data using Geographical

Information Systems (GIS) is rapidly improving this

situation. Nevertheless, an administrative commitment to

long-term environmental monitoring, data analysis, and

1

 Ecosystem Team Leader, USDA Forest Service, International Institute of

Tropical Forestry, Río Piedras, PR.
198

synthesis is required to develop adequate assessments and

verify ecological and resource management models.

Technological Change

The ability to provide safe drinking water depends on how

technology is used to: (1) measure the quantity and quality

of water, (2) run treatment plants and distribution systems,

and (3) charge consumers for the water they use and the

pollution they generate. Advances in the technology used to

accomplish these tasks are expected to improve municipal

water management. Meanwhile, other anthropogenic

activities can be expected to create new chemicals, new

pathogens, and presently unknown water-quality problems.

Increased competition for water resources and increased

emphases on instream water quality are anticipated to

increase the scrutiny of water supply and pollution management. Fortunately, recent advances in the technology used to

acquire hydrologic data are greatly improving our ability to

measure and monitor water resources. At the regional scale,

advances in climate modeling and remote sensing have

increased the ability to monitor precipitation and water

resources over large areas. At the local scale, automated

sensors and wireless communication systems are monitoring

streamflows and water withdrawals, nutrient and pollution

concentrations, stream channel morphology, and the

migration of aquatic organisms. Moreover, real-time

monitoring is currently being used to manage irrigation

systems, water supply reservoirs, flood warning systems,

and biotic migrations in rivers. Future developments are

expected in the technology to monitor pesticides, special

chemicals, and microbiological constituents. If pollution

standards based on total maximum daily watershed loads

replace standards based on average point-source discharges

(U.S. EPA 1991), we would expect improvements in the

technology for low-cost monitoring of temporal variations

in water quality.

New technology and scientific understanding can be

combined to improve the timing of land-use treatments.

Recent examples of this type of management include:

(1) the timing of streamwater withdrawal and release to

minimize impacts on the migration of aquatic organisms

(Benstead and others 1999, Bistal and Ruff 1996); (2) the

timing of fertilizer application to growth phases of agricultural crops to minimize nutrient runoff (Matson and others

1998); (3) timing the abundance of grazing with the growing

season of range and riparian vegetation (chapter 14); and

(4) the scheduling of insecticide application to the life

cycles of pests (Balogh and Walker 1992; chapter 13). The

success of these life-history-based management schemes

requires detailed knowledge of local environmental conditions; high-quality, spatially explicit monitoring; and

institutional memory of past successes and failures.

While automated data collection techniques are essential to

accurately monitoring hydrological and ecological processes,

it will remain technically and economically impossible to

monitor and treat for all contaminates, at all locations, at all

times. Furthermore, without proper analysis, the automated

collection of massive amounts of data can hinder rather than

assist management. A major challenge for watershed

management in the 21st century will be the development of

spatially explicit analytical methods and institutional

structures that can rapidly synthesize information about

environmental conditions so managers can make informed,

defensible decisions in a timely fashion. In response to this

challenge, many natural resource organizations are developing data management and environmental decision-support

systems (Lovejoy and others 1997, Spencer 1996). These

decision-support systems will eventually integrate real-time

hydrologic measurements with GIS and multiobjective

decision models. Multiobjective models have been used for

decades in the design and operation of reservoirs and water

and wastewater distribution systems because they provide a

formal and logical structure for organizing and synthesizing

scientific, environmental, and social information (Hipel

1992). Nevertheless, their adoption as to real-time management tools will be a considerable challenge. Developing and

verifying site-specific models to establish maximum daily

pollutant load allocations for specific management practices

or individual landowners will also be a major research and

management challenge (U.S. EPA 1991, 1997). Moreover,

just collecting real-time, water-quality information to be

used in complex models can be a daunting task (Bistal and

Ruff 1996).

Administrative Change

Because of high engineering and environmental costs

associated with developing new water supplies, the emphasis in water management is shifting from the development of

new sources toward the efficient and equitable use of

existing supplies (Frederick 1993, Kneese 1993). Likewise,

the costs and risks associated with transforming polluted

water into potable water are increasing the emphasis on

maintaining source water quality. In response to these shifts,

the administrative structures and organizational relationships

used to manage municipal water systems are also changing.

One reflection of this change is the number of regional

water management councils, stormwater utilities, waste

management districts, and watershed restoration groups that

have recently been established (Mann 1993, Shabman 1993,

Future Trends and Research Needs in Managing Forests and Grasslands as Drinking Water Sources
199

Taff and Senjem 1996) (see examples in appendices A, B).

These new organizations typically develop to help manage

the complexities associated with the mismatch between

natural water regions, political districts, and the geography

of water demand and wastewater generation.

While new organizational structures are developing in some

regions, many traditional water and pollution management

organizations are caught in a vicious cycle (World Bank

1993). Because of unreliable and poor-quality services,

consumers are unwilling to support increase tariffs for water

management and pollution control. Inadequate operating

funds lead to further deterioration of services by the

overextended agencies. To successfully deal with these

problems, economists have championed “user-pays” and

“market-based” approaches to water and pollution management for over 45 years (Busby 1955, Kneese 1993). For

municipal water systems, this approach has typically meant

establishing variable rate structures to promote conservation

and/or the privatization of water management services.

Several different market-based approaches have been

promoted for pollution control, including the use of effluent

charges, tradable effluent permits, tax on the use of substances that threaten water quality, and open competition for

interbasin or interregion water or pollutant transfers (Mann

1993, Taff and Senjem 1996).

While these market-based approaches have been widely

promoted by economists, water resource managers have

been less enthusiastic about their adoption (Brookshire and

Neill 1992, Taff and Senjem 1996). The principal stumbling

blocks are usually the technological and organizational

requisites for monitoring and enforcing the complex and

dynamic trade of water resources and pollutant effluent.

Recently developed data acquisition technology allows

pollution discharges and water withdrawals to be monitored

continuously. It, therefore, is generally believed that these

market-based, watershed-based approaches will increase in

the near future as water-quality regulations based on

watershed-wide maximum pollution loads are implemented.

Site-Specific Considerations

The social and environmental responses to changes that can

affect municipal watersheds are complex because different

ecosystems and processes respond at different rates and in

different magnitudes (Schimel and others 1996). Furthermore, the risk and consequences of inaccurate decisions are

not evenly allocated across the landscape or population

(Frederick 1993). Therefore, site-specific research and

monitoring are needed to develop local and ecosystemspecific understandings of the processes that effect water

quality at a particular location. For example, removal of a

road requires site-specific analysis, or the disturbance

caused by road closure may accelerate rather than reduce

erosion (Elliot and others 1996) (see chapter 9). Likewise,

the impacts of grazing also require an understanding of the

specific grazers and the local, seasonal cycle of rangeland

vegetation (see chapter 14). The response of atmospheric

deposition of nitrogen also depends on the level of nitrogen

saturation of the receiving ecosystem and seasonal variations in plant growth and nitrogen use (Fenn and others

1998). Toxic algal blooms in lakes can occur in specific

portions of some lakes and result from unique combinations

of site-specific climatic events and management operations

(James and Havens 1996) (see chapter 5, Hebgen Lake case

study). The abundance and frequency of herbicide use also

depend on site conditions and can range from never to

several times each year (Balogh and Walker 1992) (see

chapter 13). Furthermore, the width and composition of

buffer zones needed to contain certain chemicals, reduce

impacts of grazers, or maintain aquatic habitat also are

ecosystem and problem specific. Water-quality changes due

to ozone-induced stress on conifers are also closely related

to location-specific forest cover changes (Graham and

others 1991). Likewise, the impacts of accidental chemical

spills or other historical legacies that alter water quality are

site-specific and require local knowledge and institutional

memory to be properly assessed and efficiently managed.

The importance of understanding the timing of specific

environmental events and processes is an additional theme

in many of the chapters in this report. For example, the

amounts and impacts of recreation on site and water quality

are regional, seasonal, and episodic (see chapters 6, 7, and

8). The use and, therefore, the potential influence of wildlife

on water quality can also vary with seasonal and diurnal

behavior and the abundance of specific populations (see

chapters 15, 16, and 17). Grazing behavior also varies with

season, species, and the age of individuals (chapter 14).

Likewise, determining life histories and vectors of waterborne human pathogens is also essential for evaluating sitespecific risks and management options.

The length of time that a particular activity affects water

quality also varies with land use and site-specific characteristics. Sediment yields or concentrations following timber

harvesting typically decrease as a negative exponential

relationship while changes in nutrient concentrations occur

in relatively brief pulses (see chapter 10). Sediment yields

from roads typically peak in the first few years but can

remain elevated for decades, while contamination from

roadside fuel spills can last for years (see chapters 9, 11).

Likewise, the residence time of fecal contamination in

streams can be on the order of weeks to months (see chapter

2). Mining debris can acidify surface and subsurface water

for decades or longer (see chapters 18, 19).

Chapter 21
200

Developing methods and monitoring protocols to determine

and predict environmentally critical time periods is an

additional challenge that will require the interaction of

scientific information, technology, and administrative

structures. The future success of these interactions will

depend on the availability of high-quality, spatially explicit,

long-term environmental data. In many regions, assessments

of municipal water supplies will provide invaluable baseline

information for future managers.

Conclusion

The flow of water across the landscape and its collection

and distribution through a municipal water system are

complex and dynamic processes. Because of the complexities and risks involved, some basic level of water treatment

and monitoring is always necessary. However, it is technically and economically impossible to monitor and treat for

all contaminates, at all locations, at all times. Providing the

necessary levels of watershed protection, treatment, and

monitoring to sustain supplies has been, and will continue to

be, a major challenge. The dramatic improvements in U.S.

water quality that have occurred during the last few decades

clearly demonstrate the success that integrated, continued

management can have. It is hoped that this report will assist

water resource managers to successfully identify critical

problems and protect the best and restore the rest.

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Chapter 22

Synthesis

Douglas F. Ryan1

Introduction

Forest and grassland watersheds have traditionally been

relied upon as sources of drinking water with relatively little

contamination. Recent results from the National WaterQuality Assessment (U.S. Geological Survey 1999) reaffirm

that, nationwide, water from forest and grassland watersheds

is lower in many pollutants than that from watersheds

dominated by urban or agricultural land use. These findings

do not contradict the scientific evidence reviewed in this

report, which shows that many types of common land-use

practices and natural processes in forests and grasslands can

introduce contaminants into water sources. Rather, taken

together, these results indicate that although land-use

practices in forests and grasslands can introduce contaminants, the characteristics and intensity of these practices,

when applied over large areas, produce water that is cleaner

in many respects than other, more intensive land-use

practices. At the local level, forest and grassland management may cause significant problems for drinking water

sources. For example, high-intensity activities such as

logging, mining, or urban-style development in forests can

cause considerable pollution as can uncontrolled events such

as floods, landslides, or accidental chemical spills. At the

regional level, contaminants from forests and grasslands,

even where low in concentration, are part of the overall,

cumulative load of water pollution. Thus, assessing the risk

of contamination for drinking water systems with source

areas in forests or grasslands is not fundamentally different

from assessing risks in areas with other types of land uses.

Regardless of land use, assessments should be done on a

case-by-case basis, analyzing the natural processes and

human activities that can reasonably occur in the source area

to estimate the likelihood that contaminants will be transmitted to a drinking water intake.

Drinking Water Contaminants and Treatments

Contaminants of concern for drinking water have been

classified and standards for acceptable levels set by the U.S.

Environmental Protection Agency (EPA). The relationship

between specific forest and grassland best management

practices and drinking water quality is complex and was not

treated in detail in this report. Best management practices

and the protection that they provide for water quality vary

considerably from State to State and are evolving over time.

As effects on human health from contaminants in drinking

water become better understood and as new substances are

released to the environment, changes in drinking water

standards can be expected in the future. Standards for

drinking water do not apply to source water before it has

been treated to remove contaminants. Standards set under

the Clean Water Act (Public Law 80–845) that apply to

ambient water as it flows in a stream or lake are not intended to ensure that water is drinkable without treatment.

Considerable treatment may be required to purify water that

meets the ambient standard to make it comply with the

drinking water standard. For examples of forest and grassland management practices that have been proposed to

protect water quality, see U.S. EPA (1993), copies of which

can be ordered from the Web at http://www.epa.gov/

OWOW/info/PubList/publist4.html.

Drinking water treatment technology can be designed to

reduce most contaminants in source water to an acceptable

level before delivery to consumers. The cost of treatment,

however, usually increases substantially as the amount of

contamination in source water increases. Adopting appropriate land uses and management practices that do not contaminate source water has the potential to be more cost-effective

than treatment of source water that has been contaminated.

Cumulative Effects

Although different types of land use are treated separately in

this report, in an individual watershed many different land

uses affect source water quality simultaneously. Land uses

occur in complex patterns that overlap on the landscape and

change over time. Relatively few studies have examined the

1

 Staff Watershed Specialist; Wildlife, Fish, Watershed and Air Research

Staff; USDA Forest Service; Washington, DC.
203

cumulative effects of several land uses distributed over time

and space. Most studies have focused on relatively small

areas and short time periods. Some of the tools that are

needed to analyze cumulative effects at large scales have

only become available recently. More information is needed

on the interactions among multiple land uses in complex and

changing patterns because complexity is typical of watersheds in which land management is likely to affect public

drinking water sources.

Effects of Natural Processes

and Human Activities

Drinking water sources are affected by numerous natural

and human-influenced processes that occur in watersheds.

The processes by which water and contaminants move

through watersheds are relatively well known. The risks

these processes pose to source water can be severe, especially during extreme natural events such as floods and

landslides.

Water quality may be affected where water is dammed,

impounded, diverted, or augmented for a variety of human

purposes that may or may not be related to drinking water

use. These activities often alter flow rates and residence

time of water. They change turbidity, sediment storage and

transport, and oxygen content of water. Considerable

information is available about effects of these manipulations

on source water, but this information must be applied on a

case-by-case basis. Removal of old dams may pose risks for

drinking water sources downstream. Sediments and toxic

contaminants or both in the material accumulated behind the

dam can be mobilized. Few studies, however, have evaluated this risk.

Intermixed urban and wildland uses, developed administrative sites, and concentrated recreational sites share a number

of similar risks for source water. Runoff from impervious

surfaces and improperly functioning sewerage treatment

facilities can contaminate surface water, while poorly

performing septic systems and leaking underground fuel

storage tanks can pose risks for both surface and ground

water. Systems that are old, inadequately designed, and/or

poorly maintained represent the highest risks. Forms of

concentrated recreation that involve direct water contact,

such as bathing beaches, may have a high potential for

contaminating surface water but have been little studied.

Dispersed recreation activities that attract people to spend

time near water bodies, but lack developed sanitary facilities, can introduce fecal organisms, presumably including

pathogens, into surface waters. However, few studies have

been done to determine thresholds above which dispersed

use contaminate source water excessively. Many other

potential effects of dispersed recreation on drinking water

sources, such as risks associated with pets and off-road

vehicles, are poorly understood and need further research.

Roads and other utility corridors pose risks of contamination

because they concentrate many human activities. Roadside

recreation facilities are often centers of dispersed recreation.

Roads also support transport of chemicals, some of them

toxic. Chemicals may be spilled during accidents; the

highest risk of water contamination is where roads cross

streams. Utility corridors present similar risks due to

incidents such as pipeline failures or transformer fires that

may spill chemicals. Road and corridor construction,

maintenance, and use have been shown to contribute

sediments to streams because they have elevated erosion

rates and can increase the risk of landslides on unstable

terrain. Proper engineering design, construction and

maintenance of roads and utility corridors, as well as

emergency preparedness can reduce but not entirely

eliminate these risks to source waters.

Many researchers have studied the impacts on water quality

from manipulating forest vegetation for purposes such as

timber and fiber production including growing trees,

harvesting them, and reestablishing forest vegetation. The

primary contaminant to source water from these activities is

sediment associated with soil disturbance during harvesting

and regeneration and erosion from roads. Enrichment of

streamwater by nitrate after forest harvesting has been

reported in some parts of the country. Mounting evidence

suggests that this response in some regions may be explained by mobilization of long-term accumulations in

forest soils of nitrogen compounds that were deposited from

air pollution. The degree and areal extent of this effect needs

further study.

Most modern pesticides and herbicides that are currently

used on forests and grasslands are immobilized and degraded in soils to the extent that they pose little contamination risk to source water if required application precautions

are followed. Even though the use of pesticides that resist

breakdown in the environment has largely been discontinued, some of these substances may persist in the landscape

as a result of past use. They may pose problems for source

water, especially if deposits of these chemicals in soils or

sediments are mobilized by disturbance.

Prescribed fire is normally conducted under conditions

when fire severity is low and impact on source water quality

under these conditions also is low. Wildfires, however, can

be severe. When they occur on steep or erodible terrain and

are followed by intense rainfall, they can produce large

Chapter 22
204

sediment loads that pose problems for source waters. The

causes of high nitrate levels in streams after fires in areas

with high air pollution, and the effects of fire retardant

chemicals on source water need further study. The effectiveness of emergency rehabilitation practices to stabilize watersheds after severe wildfires also needs more investigation.

Land management that results in domestic or wild animals

being concentrated near surface water can contaminate

source water, but in most cases the risk of transmission to

source water is little understood. Domestic grazing animals,

such as cattle, have been shown to introduce sediment and

fecal organisms into surface waters. Streamside buffers and

other practices that reduce contact between grazing animals

and surface water can reduce contamination risks. Although

many wildlife species are known to carry pathogens,

relatively little is known about the risk that wildlife-carried

pathogens pose for source water. The few studies of wildlife

effects on water quality mainly involve ungulates, such as

elk and deer that pose risks analogous to domestic grazing

animals when wild ungulates reach high concentrations near

surface waters. Fish hatcheries can introduce fecal matter

and chemicals if these substances are flushed from hatchery

facilities to source waters. Less-intensive forms of fisheries

management, such as altering fish habitat by introducing

large woody debris and restoring habitat for anadromous

fish runs, may also have effects on source water but these

aspects have been little studied. Large concentrations of

water birds can introduce pathogens into surface water, and

measures are sometimes taken to discourage water bird

congregations near drinking water intakes.

Mining has the potential to contaminate source water with

sediments, acids, toxic metals, and other introduced chemicals from mining residues and ore processing. Many of these

effects may be abated or mitigated during active mining but

can pose long-term risks to source water if mines are not

properly decommissioned after mining ceases. Oil and gas

exploration and extraction can cause risks from spills of oil

and drilling fluids and cross-contamination of aquifers if

well casings are not properly sealed. Abandoned wells that

are not adequately capped may be used for illegal dumping

that can contaminate source water.

Implications of Scientific Uncertainty

Evaluating risks associated with managing forests and

grasslands that are sources of public drinking water requires

close collaboration among managers, scientists, and the

public. Clearly, the most current scientific findings need to

be considered in this process, but the participants in this

joint effort will need to recognize that science alone cannot

solve many management problems. An appropriate role for

science is to provide the public, in the case of governmentowned land, or the responsible party, in the case of private

land, with a better understanding of the effects of land-use

decisions on drinking water sources.

As should be apparent from this report, there are many

situations in which scientific studies provide limited or, in

cases where knowledge is scant, almost no basis for

evaluating risks of some activities to public water sources.

To cite a few examples: the risks to source water from

dispersed recreation, from the deposition of nitrogen

compounds from air pollution, or from newly discovered

pathogens might prove to be difficult to estimate with

current knowledge. When contaminant sources are suspected, but their effects on source water are highly uncertain, it is prudent to implement backup activities such as

monitoring of water quality at source water intakes. Monitoring may offer additional benefits beyond limiting the

immediate risk to human health. It can also allow landowners and land managers to learn from experience; they can

adjust future management practices on the basis of past

results and can help to reassure the public that their water

source is being protected even where predictions of land-use

effects cannot be made precisely. To deal with large, uncontrolled events with severe consequences, such as floods

or accidental chemical spills, emergency preparedness

measures should be instituted. In this way, public water

supplies can be protected by minimizing damage from such

potential catastrophes.

Gaps in scientific knowledge about the effects of forest and

grassland management on drinking water sources indicate

areas in which future research can provide large potential

benefits for land managers. Some examples emerged in the

preparation of this report. There has been very little research

on the risk that land management activities may introduce

disease organisms into source water. Research is needed on

the potential to transmit pathogens to source water from

urban-wildland intermixed development, from water contact

activities such as swimming as well as recreational pursuits

that occur in areas without developed sanitary facilities, and

from management that concentrates animals near source

water. Animals that may transmit pathogens include water

birds and other wildlife, as well as domestic animals such as

livestock and pets. Research is needed on how land management activities such as timber and fire management in areas

of high nitrogen deposition from air pollution may affect the

release of nitrates into source water. There is a need for

research on how forest and grassland management practices

affect the economic costs and benefits of providing safe

drinking water and to whom costs and benefits may accrue.

Synthesis
205

Implications for Source Water Assessments

The Forest Service and other government agencies are

developing relatively new management approaches, such as

ecosystem management and sustainability, to ensure the

long-term viability of species and human communities that

depend on public land. A key objective of these practices is

to consider a broad spectrum of values in land management.

Source water assessments are consistent with this approach

because they increase the public’s awareness of source water

protection as a value to be included in land management.

Teams that perform source water assessments should be

composed of individuals from a wide variety of disciplines

and should draw upon a wide spectrum of public opinion to

ensure that the many relevant points of view are included in

the analyses. In addition, these teams should make an effort

to integrate across disciplines because many risks to source

waters arise at the intersection of activities traditionally

considered to be the subject of separate disciplines. For

example, an impoundment constructed for the sole purpose

of storing source water could become a source of contamination if it attracted large numbers of water birds to feed or

roost on the reservoir or drew large numbers of recreationists to enjoy its shores.

Assessments should consider activities that take place in

sensitive parts of the landscape that have strong linkages to

surface water. Activities in these sensitive areas should be

examined carefully for contaminant risk, especially during

floods. Examples of areas with strong hydrological linkages

to surface water are: active stream channels or lakebeds,

riparian areas, floodplains, and areas near wetlands, springs,

and seeps. Surface water moves quickly through the

landscape. Contaminants that reach surface water can reach

drinking water intakes rapidly, with relatively little time for

transformation or removal by natural processes. When

incidents such as chemical spills contaminate surface water,

fast action may be needed to protect drinking water. Once

the source of surface contamination is contained, however,

the rapid movement of surface water tends to flush away

contaminants quickly. Where possible, potential contaminant sources should not be permitted in these areas, and

vegetated buffer strips should be retained in these zones.

Some sensitive areas may be linked by gravity to surface

water. Steep slopes and areas with unstable geomorphology

close to water bodies carry high risks. Activities that disturb

the soil surface or remove significant portions of the

vegetation in these sensitive areas may cause high erosion

rates or landslides that may increase sediment loads for

drinking water sources downslope.

Some sensitive zones may also have strong biological

linkages to surface water. Both wild and domestic grazing

animals are strongly attracted to surface water, especially in

arid areas or during droughts when forage may be concentrated near water. Water bird populations and migratory fish

can likewise be drawn from long distances to suitable

habitat in or near water. Management actions that cause high

animal concentrations in or near drinking water sources

need to be carefully examined for their potential to introduce contaminants.

Some sensitive parts of the landscape have strong linkages

to ground water that is used for drinking water sources.

Recharge areas that have high infiltration of water to

aquifers are particularly susceptible to contamination.

Ground water and its contaminants usually move very

slowly. This may allow long warning times for systems that

rely on ground water, e.g., when contaminants are detected

in ground water moving towards drinking water wells.

However, once an impurity is introduced into an aquifer,

contamination may be essentially permanent because ground

water movement may take a very long time to flush away

impurities. Exceptions exist where underground conduits

such as caves, permit rapid underground flows, making

predictions of underground contaminant movement very

uncertain. Facilities such as septic systems, leach fields,

underground fuel tanks, and dumps in sensitive aquifer

recharge areas are potential contamination sources and need

careful maintenance and monitoring. Illegal dumping of

toxics in recharge areas may go undetected until it reaches a

well, producing potentially serious consequences for health

and disruption of water supply.

Not all contaminants pose the same risks. Contaminants

such as toxic substances and pathogens that can cause

serious health problems if they are consumed in drinking

water obviously deserve a high priority in assessing risk.

Contaminants that do not pose a direct threat to health but

may make water less palatable, or may interfere with or

increase the cost of treatment such as clean suspended

sediment, color, odor, or taste should logically not receive as

high a priority as health-threatening contaminants, but the

economic cost of treating these nonhealth-threatening

contaminants should be included in analyses.

Source Water Protection as a Priority

for Land Management

In principle, providing safe drinking water to protect human

health is a high priority in our society. In practice, this

priority is often not well represented in land-use decisions.

As was suggested in chapter 4, appropriate land-use

practices that protect clean source water may be more costeffective for society as a whole than removing pollutants

Chapter 22
206

after the fact. However, decisions about land uses and their

effects on water are often made piecemeal and potential

savings often are not realized. Source water assessments can

help to forge the connection between land use and drinking

water protection by better informing land managers and the

public about that linkage. For society to fully benefit, other

mechanisms may need to be established that more closely

link the outcomes of land management to its effect on

drinking water sources. In chapter 4, an example was cited

of how a water utility provided an economic incentive for

land managers upstream to control contaminants. Many

other legal, institutional, or economic arrangements could

potentially produce similar, positive effects on source water

quality. Managers of land and of drinking water systems

should be encouraged to cooperate to protect drinking water

safety at the local level, but more fundamental and farreaching changes may be required at the policy level to

enhance incentives that encourage such cooperation and to

overcome obstacles that inhibit it.

The human values that drive economic and political

decisions affecting drinking water sources are reflections of

the values held by the general public. Over time, the values

that people place on natural resources can be expected to

change and the purposes that public lands serve will

likewise change. A challenge for land managers and the

public will be to see that, as new patterns of land uses arise

in the future, the importance of safe sources of drinking

water to protect human health is given due consideration

among the competing uses of forests and grasslands.

We hope that land managers, scientists, and the public will

draw upon the basic scientific information in this report and

apply it to their local watersheds as they participate in

source water assessments in their State. We expect that they

will also draw upon their own experience and use their best

professional judgement to decide what portions of this work

are most relevant to their particular setting within their

individual source areas. In doing so, they may find important factors and interactions at work in their particular

watersheds that we have overlooked, and will reveal future

research needs. We look forward to the results of their

analyses, because they will advance the understanding of the

relationship between land management and source water

quality. Their efforts will be an important next step toward

better protection of drinking water sources in future land

management decisions.

Literature Cited

U.S. EPA. 1993. Guidance specifying management measures for sources of

nonpoint pollution on coastal waters. EPA–840–B–92–002. Washington,

DC: U.S. Environmental Protection Agency. 836 p.

U.S. Geological Survey. 1999. The quality of our Nation’s water—

nutrients and pesticides: Circ. 1225. Denver: U.S. Geological Survey.

82 p.

Synthesis
207

Part VII:

Appendices

Cockolds Creek, ACE Basin, South Carolina. Photo by Bill Lea
209

Appendix A

City of Baltimore Municipal Reservoirs, Incorporating

Forest Management Principles and Practices

Robert J. Northrop1

Introduction

The city of Baltimore owns and operates the Loch Raven,

Prettyboy, and Liberty Reservoirs, located north and

northwest of the city, in the northern Piedmont region of

Maryland (appendix fig. A.1). They supply water to over

1.5 million people. The reservoirs are surrounded by 17,580

acres of city-owned forest that was acquired between 1880

and 1955 to ensure control of land use in critical areas

immediately adjacent to the reservoirs. Forest management

on the reservoir land dates back to 1919. Following the

clearing for Loch Raven and Prettyboy Reservoirs, a

logging and sawmill crew was retained for forestry work,

and the first professional forester was hired. This forest

management program was undertaken to harvest and sell

forest products while protecting the reservoir. Revenues

were used for watershed enhancements, and lumber was

used by the Department of Public Works in Baltimore.

In recent years, the reservoir land has also been valued as a

core area for the conservation of regional biodiversity and

for dispersed outdoor recreation. In 1989, concerns about

timber harvesting, uncontrolled access, and a rapid increase

in recreational use convinced the city to reevaluate its

management practices. At the same time, the public agencies

responsible for Maryland’s Source Water Protection

Assessment [Safe Drinking Water Act Amendments of 1996

(Public Law 104–182)] were expressing concern over the

eutrophic nature of the three reservoirs and their loss of

storage capacity due to sedimentation.

The watersheds, which are the primary sources of water

for the reservoirs, are in Baltimore, Harford, and Carroll

Counties in Maryland, as well as York County, PA.

City-owned land makes up only an average of 7 percent of

the total area of the watersheds draining into each reservoir.

These source water drainages are part of the urbanizing and

expanding Baltimore-Washington metropolitan area, which

is the fourth largest in the United States. The Prettyboy and

Liberty basins, however, are still rural in character with

agricultural use predominant. Preserving the quality of

the water that flows into the reservoirs requires careful

control of sediment, as well as point- and nonpoint-source

pollutants.

Our watershed management strategy is seen as vitally

important to the continued efficient and economical provision of safe drinking water for the region’s residents by all

Federal, State, and local agencies. Since forest management

can influence water quantity, as well as quality, by filtering

and sequestering various forms of soluble and solid pollutants coming from adjacent land uses, it is recognized as a

key component of the management of these watersheds.

Private forest landowners are enhancing water quality by

applying several forest conservation principles on their land.

They are restoring forest wetlands and riparian forests and

are using silvicultural practices to maintain forest vigor.

Studies

The Maryland Department of Natural Resources (MDDNR), Forest Service, has entered into an agreement with

the city of Baltimore to develop a comprehensive Forest

Resource Conservation Plan for the 17,580 acres of land

surrounding the Loch Raven, Prettyboy, and Liberty

Reservoirs. Through a cooperative agreement with the U.S.

Department of Agriculture, Forest Service, and the use of its

NED–12 Decision Support Software, a detailed forest stand

level analysis incorporating forest patch methodology will

be conducted. Additional data will be collected on wildlife

habitat composition and structure, and on the quality of

water in first- and second-order streams. A separate recreational use survey will be conducted through contract with a

regional university.

1

 Regional Watershed Forester, Maryland Department of Natural

Resources–Forest Service, North East, MD.

2

 A prescription design system that incorporates management goals for

multiple objectives, analyzes current forest conditions, produces

ecommendations for management alternatives, and predicts future

conditions under different alternatives. This system assists in evaluating

silvicultural decisions at a project level using landscape-scale factors.
210

Appendix figure A.1—Watersheds supplying Baltimore, MD, with water.

Goals for conservation were set through a series of 20 public

meetings conducted by the city of Baltimore’s Department

of Public Works during 1991. These goals included:

1. The protection and enhancement of water quality.

2. The maintenance and restoration of regional biological

diversity within the public lands surrounding the

reservoirs.

3. The management of woodlands to maximize forest

habitat value.

4. Providing recreational opportunities compatible with the

above objectives.

Concurrently, the MD-DNR Forest Service has also begun

work with the Baltimore Metropolitan Council of Governments and the Gunpowder Watershed Project. They have a

U.S. Environmental Protection Agency (EPA) small

watershed grant project, where Federal, State, and local staff

work to develop cooperative and collaborative strategies to

address various environmental issues, including source

water protection, in a holistic fashion at the watershed level.

Background data and information on the Loch Raven and

Prettyboy Reservoir drainage basins are being supplied

through the Maryland Department of the Environment’s

Source Water Assessment Program (Safe Drinking Water

Act, sec. 1453). Background data and information on the

Liberty Reservoir drainage basin are being supplied through

the Department of Natural Resources’ Unified Watershed

Assessment, as part of the State’s Clean Water Action Plan.

Both studies used land-use loading coefficients to estimate

the pounds of nutrients and sediment typically produced for

a classified land use. Composite storm event samples and

baseflow were collected at various sites in the Prettyboy and

Loch Raven drainage basins but not the Liberty basin.

Preliminary reports from the modeling exercises indicate

that there is a statistically significant increasing trend in

nitrate concentrations reaching the Loch Raven Reservoir,

with highest concentrations in baseflow, indicating historical

ground water contamination. These reports are also indicating that nutrient contamination is widespread throughout the

Prettyboy and Liberty drainage basins as well, with current

levels in the same range as at the Loch Raven basin.

Source water protection strategies being developed by these

groups and associated State agencies highlight the need to

conserve the existing forest in a healthy and vigorous

condition. Forest wetlands and riparian forests need to be

restored for their functional ability to filter sediment and

other suspended solids, sequester pollutants in woody tissue,

and promote denitrification. Forest wetland and riparian

forest restoration activities within the three drainage basins

will be targeted to specific sites that provide the best

opportunity to intercept ground water and overland flows

before they reach the receiving streams. Using a Geographic

Information System (ArcView) and data layers from various

Federal, State, and local agencies, the MD-DNR has

developed a method that locates and ranks potential restoration sites. Potential forest wetland restoration sites are

located by identifying hydric soils that lack natural vegetative cover. This system also locates potential riparian forest

restoration sites by identifying stream segments that lack

forest cover and assessing their potential to improve water

quality. A weighed ranking is assigned based upon the

nutrient loading potential of adjacent land uses, the size of

the ownership parcel, and stream order (lower order streams

receive higher ranking).

Interest in the management of the city-owned and surrounding forest is keen. Public support is critical to the plan’s

successful implementation. The Friends of the Watershed,

an existing city-sponsored citizen’s advisory group, will be

invited to review data sets that are being collected, as well

as the proposed analysis. They will also be asked to assist in

the identification of public meeting sites and the context for

stakeholder involvement.

City of Baltimore Municipal Reservoirs, Incorporating Forest Management Principles and Practices

Loch

Raven Pretty Boy

Liberty
211

Anticipated Results

Through the analysis of forest resources at multiple scales

(unit to ecoregion) and timeframes, city-owned tracts will be

evaluated to determine their potential to:

1. Serve as buffers to adjacent land uses.

2. Support an increasing desire on behalf of the growing

urban population for outdoor recreation.

3. Assist in the conservation of biological diversity at the

regional scale.

The comprehensive forest conservation plan will provide

explicit management recommendations, allowing the city to

plan and organize its conservation activities in the most

Appendix A

efficient and effective manner. The deliberate and comprehensive involvement of interested citizens and community

associations will lead to the public consensus the city needs

to once again feel comfortable in actively managing its

properties for the multiple values consistent with its stated

goals.

Finally, the plan will offer forest management guidance to

Federal, State, and local agencies concerned with the

continued decline in the region’s forest land base. This

decline has been compounded by the cumulative impacts of

pollution, fragmentation, and habitat loss. The plan will

support and clarify the functions of forest resources as

integral to the long-term sustainability of local watersheds

and lead to the incorporation of forest management techniques into watershed strategies concerned with water

quality.
212

Appendix B

Managing the Shift from Water Yield to Water Quality

on Boston’s Water Supply Watersheds

Thom Kyker-Snowman1

Boston’s drinking water derives from surface reservoirs

within three major watersheds: the Quabbin, Ware River,

and Wachusett (appendix fig. B.1). These watersheds total in

excess of 225,000 acres [90 000 hectares (ha)]; about 40

percent is under the care and control of the Metropolitan

District Commission’s Division of Watershed Management

(MDC-DWM). This system supplies approximately 250

million gallons [900 million liters (L)] daily to accommodate the drinking water demands of 2.5 million people,

about 40 percent of the population of the Commonwealth.

The water is currently treated (chlorine and chloramines for

disinfection, fluoride to promote healthy teeth, and soda ash

and carbon dioxide to prevent corrosion of pipes), but not

filtered. The objective to avoid the costs and the many other

ramifications of filtration is at the center of current management decision-making for these watersheds. This objective

represents a dramatic shift from the focus on water quantity,

which has dominated the history of Boston’s water supply.

Since the settlement of Boston, its citizens continued to

look west to meet the increasing demand for water. In 1795,

the Aqueduct Corporation was created to tap Jamaica Pond

in Roxbury to supply the 20,000 Boston inhabitants. In

1848, Lake Cochituate was added, and in short order from

1870–80, the Sudbury River and Framingham Reservoirs

came on line. By 1895, Boston’s population exceeded

500,000 and the metropolitan area exceeded a million. The

Wachusett Reservoir, the largest reservoir in the World at

the time, was built by 1908 and added a 65 billion gallon

(234 billion L) capacity to the system. This was still not

enough to keep up with the growing demand. Then Boston

tapped the Ware River with an aqueduct to the Wachusett

Reservoir, and finally constructed the 412-billion-gallon

(1500-billion-L) Quabbin Reservoir within the Swift River

Valley.

Despite these efforts, water quantity persisted as a concern.

In 1967, just 20 years after the Quabbin Reservoir filled to

capacity, a severe drought lowered the reservoir to 45

percent of its capacity and skeptics worried it would ever fill

again. Although Quabbin Reservoir filled to capacity again

by 1976, water demands were exceeding the safe yield from

the system (300 million gallons per day or 1.1 billion L per

day) by almost 50 million gallons (180 million L) per day.

After lengthy debates about augmenting supplies by

diverting the Connecticut or Millers River to the reservoirs,

the MDC-DWM was mandated to address the situation by

increasing water yield from its lands. The primary approach

was to clearcut 2,000 acres (800 ha) of red pine (Pinus

resinosa Ait.) plantations and to convert them to grass fields,

which was estimated to provide an additional 300 million

gallons (1.1 billion L) of water annually. After a 1989

drought dropped Quabbin Reservoir to a 17-year low,

authorities declared a water emergency. Water conservation

efforts (spurred in part by raising water rates), and an

aggressive leak detection and repair program have dramatically lowered water consumption. Today, the daily draw on

the system is 50 million gallons (180 million L) below its

safe yield.

In addition to converting pine plantations to grass,

MDC-DWM postponed management of an inflated deer

population, because it was thought that deer browsing the

understory could increase in water yield. The deer population in the Quabbin watershed had grown to nearly 48 to 80

deer per square mile (19 to 31 deer per square kilometer)

(6 to 8 times the statewide average), under the hunting

restrictions on MDC-DWM lands. Early forest management

plans had acknowledged the impact of this population on the

understory. The emphasis on water yield made it easier to

choose to avoid the difficult politics associated with starting

a deer management program, especially following 50 years

of hunting prohibition.

Changes in drinking water laws and regulations have

dramatically altered the approach to managing natural

resources on the watersheds whose waters are unfiltered

surface supplies. The Federal Safe Drinking Water Act

(SDWA) became law in 1974, and set national standards for

maximum contaminant levels and treatment techniques. 1

 Natural Resources Specialist, Massachusetts Metropolitan District

Commission, Division of Watershed Management, Belchertown, MA.
213

Appendix figure B.1—Location of Quabbin Reservoir and water supply system for Boston, MA.

Amendments to the SDWA in 1986 established a priority for

using filtration as a dominant treatment technique. The EPA

addressed this priority through the Surface Water Treatment

Rule of 1989 (SWTR), which essentially required that all

surface water supplies be filtered unless a supply could pass

a rigorous test allowing it to qualify for a waiver from

filtration. The SWTR established disinfection and monitoring requirements and set new limits for pathogens and

turbidity, which indicate the success or failure of either

artificial or natural filtration.

It has been estimated that the construction costs alone for a

filtration plant for Boston’s water supply would exceed $200

million. This alone is a strong incentive to maintain a

waiver, but perhaps more important is the threat of losing

the mandate for watershed protection, should filtration

become a reality. The MDC-DWM currently owns and

controls 64 percent of the Quabbin watershed (appendix

fig. B.1), and this control is a critical argument in favor of

relying on natural filtration. If artificial filtration were

installed, it is worth wondering if the budget required to

manage MDC-DWM lands and to pay tax substitutes to the

local towns would persist. Similarly, recreation is carefully

limited on these watershed lands, and it would be increasingly difficult to resist these pressures in the absence of a

requirement for natural filtration of Boston’s drinking water.

The combination of reduced pressure to increase yields and

of the increasing desire to avoid filtration have shifted the

management focus in the Quabbin watershed away from

water production and sharply toward water-quality protection. From the natural resources perspective, this meant

demonstrating that wildlife and forest are being managed to

avoid degrading and, if possible, improving the natural

filtration process. Two major wildlife issues were met

squarely along these lines: water birds (in particular, gulls

and geese) and white-tailed deer. Seagulls threaten the

maintenance of water-quality standards when they spend

their days feeding in landfills and returning to roost by the

thousands on open surface water supplies, transporting

pathogens that can threaten human health. The MDC-DWM

Appendix B

Quabbin Reservoir Watershed:

a section of the MDC/MWRA

water supply system

Connecticutt R.

5 kilometers

0 5 miles

0

Quabbin

Reservoir

Watershed

Ware River

Watershed

Wachusett

Reservoir

Watershed Western Mass. communities

Eastern and

central Mass.

communities

Massachusetts Executive Office of Environmental Affairs - 1985

Massachusetts

Geographic

Information

System

MDC land

Quabbin Reservoir

The Commonwealth of Massachusetts

preserving the past...

protecting the future

Watershed Management

Boston

Harbor Aqueduct

NORTH
214

has devised an elaborate gull-harassment program that deals

with the problem by moving roosting birds far from the

water supply intakes.

The browsing by the high populations of white-tailed deer

consumes the forest understory and threatens the regeneration of forest cover if it is lost to natural or human disturbance. The threat of major overstory losses associated with

catastrophic hurricanes can recur in New England every 100

to 150 years; the most recent was in 1938. A model to

predict hurricane damage was developed by Harvard

University on their forest in Petersham, MA, immediately

adjacent to the Quabbin Reservation. This model predicted

in 1992 that 50 to 75 percent of the conifers and 25 to 75

percent of the hardwoods in the Quabbin watershed would

be damaged by such a storm. During the writing of the most

recent Quabbin land management plan, it was decided that

an even-aged, relatively mature forest with greatly impaired

regenerative capacity was incompatible with the desire to

maintain predictable long-term natural filtration of the

drinking water supply.

The first step in reversing this untenable condition was to

reduce the impact of deer, primarily through controlled

hunting. The MDC-DWM engaged in a lengthy, multiyear

public campaign for support of this idea, which overcame

opposition including a Federal lawsuit filed by an animalrights organization. This suit claimed that there was a

probability that an unrecovered deer, wounded by a lead

slug, would die and be fed upon by a bald eagle, which

might in turn ingest lead from the wound and die as a result.

At that time, this would have constituted an illegal taking of

a Federally protected, endangered species. The plan to

reinstitute hunting, in order to protect the drinking water

supply, persisted through this debate, and the first hunt in 50

years was conducted in 1991. Hunting has continued since

then, and regeneration of both trees and other understory

plants has been dramatic as a result. Wildflowers like

trillium (Trillum spp.) and marsh-marigold (Caltha spp.)

that were not found before hunting have reappeared after a

long absence.

In addition, plans called for diversifying both the age and

the species structure of the watershed forest cover. This

objective calls for maintaining an understory as the reserve

forest; a midstory for its rapid nutrient uptake; and an

overstory for its regulation of organic decomposition, its

provision of seed, and the water infiltration and retention

function of its deep root system. These canopy layers are to

be balanced, in an uneven-aged silvicultural approach,

throughout the managed forest surrounding MDC-DWM

reservoirs. This deliberate restructuring is accomplished

through commercial harvesting using primarily group

selection and irregular shelterwood approaches. The

drinking water supply context mandates state-of-the-art best

management practices, including a requirement that all

equipment be supplied with a spill kit for potential oil leaks

and strict restrictions on ground pressures allowed on

sensitive land.

The working hypothesis of this approach is that frequent,

endogenous disturbance of the scale of group-selection

silviculture will lessen the amplitude of the disturbance

wave represented by infrequent, exogenous disturbances,

such as catastrophic hurricanes. The MDC-DWM made the

commitment that any short-term negative effects of timber

harvesting would not exceed the long-term benefits to

drinking water derived from this deliberate forest structuring. While the large volume of Quabbin Reservoir dilutes

differences in tributary water quality, the no-net-negative

policy will require intensive monitoring at the tributary

level, especially during storm events and spring runoff. This

monitoring effort has recently begun at Quabbin and will

hopefully quantify the effects of incorporating large,

infrequent disturbances into management planning for

unfiltered surface supplies of drinking water.

Managing the Shift from Water Yield to Water Quality on Boston’s Water Supply Watersheds
215

Appendix C

Cumulative Impacts of Land Use on Water Quality

in a Southern Appalachian Watershed1

Wayne T. Swank and Paul V. Bolstad2

Introduction

Water-quality variables were sampled over 109 weeks along

Coweeta Creek, a fifth-order stream located in the Appalachian Mountains of western North Carolina. The purpose of

the study was to observe any changes in water quality over a

range of flow conditions with concomitant downstream

changes in the mix of land uses. Variables sampled include

pH, bicarbonate (HCO3

-1), conductivity, nitrate nitrogen

(NO3-N), ammonium nitrogen (NH4+-N), phosphate

phosphorus (PO4

3—P), chloride (Cl-1), sodium (Na+

),

potassium (K+), calcium (Ca2+), magnesium (Mg2+), sulfate

(SO4

-2), silica (SiO2

), turbidity, temperature, dissolved

oxygen, total and fecal coliform, and fecal streptococcus.

Landcover and land use or both were interpreted from

1:20,000 aerial photographs and entered in a Geographic

Information System, along with information on total and

paved road length, building location and density, catchment

boundaries, hydrography, and slope. Linear regressions were

performed to related basin and near stream landscape

variables to water quality.

Five water-quality monitoring stations were located over

5.4 miles (8.7 kilometers) of Coweeta Creek (appendix fig.

C.1). Along Coweeta Creek, stream size and permanent

landscape alteration increases, e.g., conversion of forest to

agriculture and increases in road density, from lower to

higher station numbers (appendix table C.1). Sites were

selected to encompass incremental additions and a variety of

land uses. Most of the area above station 1 was covered with

mature, deciduous forest, and paved road density was low,

while unpaved road density was relatively high. Downstream stations were selected to encompass additional landuse features such as residences along the stream, grazing

and other agricultural practices, plus additional roads.

Stations 2 through 4 were characterized by a 6- to 20-foot

(2- to 6-meter) wide riparian shrub strip [chiefly alder,

(Alnus spp.), bramble (Rubus spp.), and willow Salix spp.)]

with a mix of pastures, homesites, and farmland beyond the

riparian strip. Station 5 was in a low-density suburban mix,

with mown grass to the stream edge.

Streamwater samples were collected during baseflow and

stormflow periods. During baseflow, grab samples were

collected in 1-liter bottles from the free-flowing section of

the stream. Sampling was initiated the first week of June

1991 and was conducted twice weekly through August.

Thereafter, baseflow sampling was conducted approximately weekly through the first week of November 1993.

During selected storm events, two different sampling

methods were used. Grab samples were taken on the rising

limb of the hydrograph, near peak flow, and on the hydrograph recession. Some storm events were also sampled

using a time-proportional automated sampler, which was

activated near storm onset.

1

 This example is excerpted from a paper by Bolstad and Swank (1997).

2

 Retired Project Leader, USDA Forest Service, Southern Research Station,

Otto, NC; and Professor of Forestry, University of Minnesota, Minneapolis,

MN, respectively.

Appendix figure C.1—Watershed boundary and stream sampling locations

in the Coweeta Creek Watershed in western North Carolina. Stations 1

through 5 are arranged down the stream gradient on Coweeta Creek. Firstorder streams are not shown.

Coweeta Creek

North

Station

Stream

Watershed

0 2000 meters

5

4

2 3

1
216

Stormflow Water Quality

Conductivity, NO3-N, HCO3

-

, Cl-, K+, Na +, Ca2+, Mg2+,

SiO2, turbidity, temperature, and total coliform often showed

cumulative increases downstream. Two patterns were

obvious in comparing stormflow and baseflow data. First,

mean values for most variables at most stations were higher

during stormflow. These increases range from slight and

nonsignificant to quite large (turbidity, appendix fig. C.2).

Baseflow Water Quality

Water quality was good during baseflow conditions over the

3-year study period. Concentrations of most solutes averaged <1 milligram per liter, typical of stream chemistry for

lightly disturbed forest watersheds in the Southern Appalachians. NO3

-N, NH4+-N, and PO4

3—P were very low,

indicating the absence of point sources of inorganic solutes

into the stream. Turbidity during baseflow was generally

low, typical for the Southern Appalachians (appendix fig.

C.2), averaging <6 nephalometric turbidity units for all

stations. Mean counts of total fecal coliform and fecal

streptococci at station 1 were typical of mean values

reported for other streams draining relatively undisturbed

forested watersheds in western North Carolina. Several

variables showed distinct downstream increases. Cation

concentrations, SiO2, HCO3

-1, SO4-2, Cl-1, conductivity,

turbidity, and temperature generally increased downstream

from station 1 to 5.

Mean baseflow levels for total coliform, fecal coliform,

and streptococci counts increased from threefold to eightfold downstream (appendix table C.2). Thus, there is a

cumulative increase in bacteria populations, indicating

additive sources downstream. The transport of these bacteria

is probably primarily through the soil or direct input by

warm-blooded vertebrates, e.g., raccoons, livestock, since

base-flow samples represent periods when there is little or

no overland flow input from adjacent lands. Appendix figure C.2—Mean and standard error (bars) for turbidity,

plotted against building density for each sampling condition (baseflow

and stormflow).

Cumulative Impacts of Land Use on Water Quality in a Southern Appalachian Watershed

Appendix table C.1—Summary data for the catchments above five sampling stations along Coweeta Creek

in western North Carolina

Sampling station number

Characteristics upstream Upstream Downstream

of sample station Units 1 2 3 4 5

Total area Ha 1605 1798 3099 4163 4456

Forest area Ha 1600 1782 2986 3904 4113

Agricultural area Ha 4 13 89 155 192

Urban/suburban area Ha 1 3 24 104 151

Total road length Km 39.8 45.2 80.8 106.8 122.6

Unpaved road length Km 38.6 43.9 73.4 96.4 106.5

Total road density Km/km2 2.49 2.51 2.61 2.60 2.75

Unpaved road density Km/km2 2.41 2.44 2.37 2.33 2.39

Structures/area No./100 ha .37 3.06 5.36 6.01 9.23

Building density (no./100 ha)

Turbidity (NTU)

Baseflow data

Stormflow

data

Upstream Downstream

0 2 4 6 8 10 12 14

50

40

30

20

10

0
217

Bacteria levels were among the most responsive waterquality variables during storm events, although patterns

were highly variable among storms and among seasons.

Total coliform, fecal coliform, and fecal streptococci

typically increased twofold to threefold during storm events

compared to baseflow populations. The source of these large

downstream increases in bacteria may be attributed to

observed overland flow from adjacent lands directly into

streams, disturbance of bottom sediments, and streambank

flushing (appendix table C.3, appendix fig. C.3).

Conclusions

In summary, this work identifies consistent, cumulative

downstream changes in Coweeta Creek concomitant with

downstream changes in land use. Furthermore, this work

indicates consistently higher downstream changes during

stormflow when compared to baseflow conditions, suggesting cumulative impacts due to landscape alteration, as tested

here, are much greater during stormflow events.

Literature Cited

Bolstad, P.V.; Swank, W.T. 1997. Cumulative impacts of landuse on water

quality in a Southern Appalachian watershed. Journal of the American

Water Resources Association: 519–533. Vol. 33, no. 3.

Appendix C

Appendix figure C.3—Mean and standard error (bars)

for (A) total coliform and (B) fecal streptococcus, plotted

against building density for each sampling condition

(baseflow and stormflow). (Building density increases

downstream).

Appendix table C.2—Summary water-quality data from

baseflow grab samples (means) at each of the five sampling

stations along Coweeta Creek in western North Carolina

Station number

Variable 1 2345

- - - - - - - - - - - Per 100 milliliters - - - - - - - - - - -

Total coliform 9,470 13,660 40,040 30,740 52,140

Fecal coliform 200 340 460 1,130 840

Fecal streptococcus 710 1,310 2,180 1,590 1,840

Appendix table C.3—Summary water-quality data

from stormflow samples (means) at each of the five

sampling stations along Coweeta Creek in western

North Carolina

Station number

Variable 1 2 3 4 5

- - - - - - - - - - Per 100 milliliters - - - - - - - - - -

Total coliform 18,790 34,640 NA 77,160 98,390

Fecal coliform 880 130 NA 970 1,260

Fecal streptococcus 450 8,710 NA 3,260 4,190

NA = Not available.

Building density (no./100 ha)

1000 colonies/100 mL

0 2 4 6 8 10 12 14

160

140

120

100

80

60

40

20

0

Total coliform by building density

Baseflow

Stormflow

(A)

Building density (no./100 ha)

Colonies/100 mL

0 2 4 6 8 10 12 14

7000

6000

5000

4000

3000

2000

1000

0

Fecal streptococcus by building density

Baseflow

Stormflow

(B)
218

Appendix D

Protozoan Pathogens Giardia and Cryptosporidium

David Stern1

Introduction

Two pathogens carried by wildlife, Giardia spp. and

Cryptosporidium spp., are of great interest in drinking water.

Giardia cysts and Cryptosporidium oocysts are parasitic

protozoans. They are active and reproduce within their hosts

and encyst to survive in the environment during transmission between hosts. Many species of wildlife have been

found to be hosts for these parasites (appendix table D.1).

These organisms are significant sources of gastrointestinal

illness (Jokipii and others 1983, Kenney 1994, Moore and

others 1994). The risk posed by these parasites is believed to

be significant. As little as one cyst may be able to cause

infection (Medema and others 1995, Rose and Gerba 1991,

Rose and others 1991). What is more, these organisms are

resistant to disinfection (Campbell and others 1982, Clark

and Regli 1993, Craun 1981, Haas and Heller 1990, Hoff

and Rubin 1987, Jarroll and others 1981, Kong and others

1988, Quinn and Betts 1993, Rice 1981, Rubin and others

1989). Although there are medications that are effective in

treating giardiasis, currently there are no drugs available to

treat infections caused by Cryptosporidium.

Federal Regulation

In 1986, the U.S. Congress recognized the threat posed by

these protozoan parasites and revised the Safe Drinking

Water Act to begin to address this concern. The Surface

Water Treatment Rule for the Safe Drinking Water Act

(SDWA), promulgated in 1991 (U.S. Environmental

Protection Agency 1996), requires all surface water supplies

to be filtered prior to distribution to the public, unless it can

be demonstrated that a certain level of purity exists and can

be maintained. The Surface Water Treatment Rule emphasizes water supply filtration because disinfecting by chlorination does not eliminate the threat posed by Giardia and

Cryptosporidium (Campbell and others 1982, Clark and

others 1989). Accordingly, prevention or filtration is

recommended as the response to this threat. In its continuance of concern for the threat posed by Giardia and

Cryptosporidium, the Federal Government enacted 1996

amendments to the SDWA to fund additional watershed

research on these organisms.

Giardia

To survive in the environment, Giardia encysts itself into a

resistant form. Giardia cysts are 5 to 15 microns in size and

oblong in shape. Early research on this parasitic protozoan

identified it on the basis of median body morphology and

the host it was found in. Accordingly, Giardia species are G.

muris, G. agilis, and G. duodenalis and are usually found in

rodents, frogs, and warm-blooded vertebrates, respectively.

This early nomenclature was due to the assumption that

Giardia was highly host-specific. More recent research has

shown that Giardia can cross-infect different species of

hosts (Meyer and Jarroll 1980).

Identification of Giardia as a waterborne parasite for

humans was first reported in the 1940’s during a study of a

disease outbreak in an apartment building in Tokyo, Japan

(Davis 1948). Giardia has more recently been reported as

the most frequently identified parasite responsible for

disease outbreaks in surface water supplies in North

America (Craun 1984). A significant portion of the literature

has reported on the occurrence, disinfection, and treatment

of Giardia cysts.

The life cycle for Giardia has been described by Meyer and

Jarroll (1980). Giardia is monoxenous, which means that all

of its life stages occur in one host. These stages include an

inactive cyst form that is capable of resisting environmental

stresses and a free-living form known as a trophozoite. The

trophozoite has a ventral sucker disk that attaches to the

intestinal wall to obtain subsidence. The life cycle consists

of: (1) a host ingesting the cyst, (2) excystation (emergence

of the trophozoite out of the cyst) occurring in the small

intestine after the cyst has been subjected to the digestive

environment, (3) the released trophozoite attaching to the

intestinal wall where it feeds and reproduces by binary

fission, and (4) some of the reproduced trophozoites encyst

within the intestine and the resultant cyst is excreted in the

infected animal’s feces to be transmitted to other hosts.

Cryptosporidium

Cryptosporidium was first identified by Tyzzer (1910) over

90 years ago as a parasite of the common mouse. Its 1

 Pathogen Program Supervisor, New York City, Department of Environmental Protection, Valhalla, NY.
219

Appendix table D.1—Species reported as hosts to protozoans Giardia and Cryptosporidium

Species Common name Parasites hosted

Pisces

Cyprinus carpio Carp Cryptosporidium

Sciaenops ocellatus Red drum Cryptosporidium

Plecostomus spp. Catfish Cryptosporidium

Salmo trutta Brown trout Cryptosporidium

Oncorhynchus mykiss Rainbow trout Cryptosporidium

Amphibia

Ceratophrys ornata Bell’s horned frog Cryptosporidium

Bufo americanus American toad Cryptosporidium

B. regularis Common toad Giardia

Rana pipiens Leopard frog Giardia

R. clamitans Green frog Giardia

Reptilia

Chelonia mydas Green turtle Cryptosporidium

Geochelone elegans Star tortoise Cryptosporidium

G. carbonaria Red-footed tortoise Cryptosporidium

Squamata Lacertilia (lizards)

Agama aculeata Kalahari spiny agama Cryptosporidium

A. planiceps Damara rock agama Cryptosporidium

Chameleo c. senegalensis Chamelon Cryptosporidium

Chamaeleo pardalis Panther chamelon Cryptosporidium

Chlamydosaurus kingi Frilled lizard Cryptosporidium

Lacerta lepida Ocellated lacerta Cryptosporidium

Chondrodactylus angulifer Sand gecko Cryptosporidium

Serpentes (snakes)

Crotalus durissus culminatus Rattlesnake Cryptosporidium

Sistrurus miliarius Pygmy rattlesnake Cryptosporidium

Lampropeltis getulus holbrooki Say’s kingsnake Giardia

Elaphe subocularis Trans-Pecos rat snake Cryptosporidium

E. o. obsoleta Black rat snake Cryptosporidium

E. o. quadrivittata Yellow rat snake Cryptosporidium

E. o. lindheimeri Texas rat snake Cryptosporidium

E. guttata Corn snake Cryptosporidium

E. v. vulpina Western fox snake Cryptosporidium

Gonysoma oxycephala Red-tailed green rat snake Cryptosporidium

Pituophis melanoleucus Black pine snake Cryptosporidium

P. melanoleucus catenifer Gopher snake Cryptosporidium

Drymarchon corais couperi Eastern indigo snake Cryptosporidium

Lampropeltis zonata pulchura San Diego mountain snake Cryptosporidium

L. triangulum Various subspecies Cryptosporidium

Nerodia h. harteri Brazos water snake Cryptosporidium

N. r. rhombifera Diamondback water snake Cryptosporidium

Boiga dendrophila Mangrove snake Cryptosporidium

C. horridus Timber rattlesnake Cryptosporidium

C. atrioca udatus Canebrake rattlesnake Cryptosporidium

C. l. lepidus Rock rattlesnake Cryptosporidium

continued

Appendix D
220

Protozoan Pathogens Giardia and Cryptosporidium

Appendix table D.1—Species reported as hosts to protozoans Giardia and Cryptosporidium (continued)

Species Common name Parasites hosted

Aves

Anseriformes

Branta canadensis Canada goose Cryptosporidium, Giardia

Anser anser Domestic goose Cryptosporidium

Cygnus spp. Tundra swan Cryptosporidium

C. olor Mute swan Cryptosporidium

Aix sponsa Wood duck Cryptosporidium

Anas platyrrhynchos Mallard duck Cryptosporidium

Mergus merganser Common merganser Cryptosporidium

Columbiformes

Columba livia Pigeon Cryptosporidium

Galliformes

Gallus gallus Chicken Cryptosporidium

Meleagris gallopavo Turkey Cryptosporidium

Coturnix coturnix Common quail Cryptosporidium

Colinus virginianus Bobwhite quail Cryptosporidium

Phasianus colchicus Ring-necked pheasant Cryptosporidium

Pavo cristatus Peafowl Cryptosporidium

Perdix perdix Grey partridge Cryptosporidium

Alectoris graeca Chuckar partridge Cryptosporidium

Numida meleagris Guinea fowl Cryptosporidium

Charadriiformes

Larus ridibundus Black-headed gull Cryptosporidium

L. argentatus Herring gull Cryptosporidium

L. delawarensis Ring-billed gull Cryptosporidium

Recurvirostra avosetta Avocet Giardia

Threskiornis spinicollis Straw-necked ibis Giardia

Passeriformes

Poephila cincta Black-throated finch Cryptosporidium

Lonchura cucullata Bronze mannikin finch, Cryptosporidium

 red cheek finch

Passer domesticus House sparrow Giardia

Zonotrichia georgiana Swamp sparrow Giardia

Sturnella neglecta Western meadowlark Giardia

Lanius collurio Red-backed shrike Giardia

Ciconiiformes

Ardea herodias Great blue heron Giardia

A. cinerea Gray heron Giardia

A. cocoi Cocoi heron Giardia

Egretta alba Great egret Giardia

E. caerulea Little blue heron Giardia

Nycticorax nyctocorax Black-crowned night-heron Giardia

N. naevius Night-heron Giardia

Butorides virescens Green-backed heron Giardia

Egretta intermedia Intermediate egret Giardia

Bubulcus ibis Cattle egret Giardia

Botaurus lentiginosus American bittern Giardia

Ixobrychus minutus Little bittern Giardia

Plegadis falcinellus Glossy ibis Giardia

continued
221

Appendix table D.1—Species reported as hosts to protozoans Giardia and Cryptosporidium (continued)

Species Common name Parasites hosted

Aves (continued)

Falconiformes

Cathartes aura Turkey vulture Giardia

Elanus caeruleus Black-winged kite Giardia

Mammalia

Marsupialia

Didelphis virginiana Virginia opossum Cryptosporidium, Giardia

Pseudocheirus peregrinus Possum Giardia

Insectivora

Sorex spp. Shrew Giardia

Blarina brevicauda Short-tailed shrew Cryptosporidium, Giardia

S. cinereus Masked shrew Cryptosporidium

Parascalops breweri Hairy-tailed mole Cryptosporidium

Myotis lucifugus Little brown bat Cryptosporidium

Eptesicus fuscus Big brown bat Cryptosporidium, Giardia

Lagomorpha

Sylvilagus floridanus Eastern cottontail Cryptosporidium, Giardia

Rodentia

Ondatra zibethica Common muskrat Cryptosporidium, Giardia

Microtus agrestis Field vole Cryptosporidium

M. chrotorrhinus Rock vole Giardia

M. pennsylvanicus Meadow vole Cryptosporidium, Giardia

M. pinetorum Pine vole Giardia

M. longicaudus Long-tailed vole Giardia

M. ochrogaster Prairie vole Giardia

M. californicus Meadow vole Giardia

M. richardsoni Water vole Giardia

Clethrionomys glareolus Bank vole Cryptosporidium

C. glareolus skomerensis Skomer bank vole Cryptosporidium

Apodemus sylvaticus Wood mouse Cryptosporidium

Rattus rattus Roof rat or ship rat Cryptosporidium, Giardia

Sigmodon hispidus Cotton rat Cryptosporidium

Erithizon dorsatum Porcupine Cryptosporidium, Giardia

Mus musculus House mouse Cryptosporidium, Giardia

Zapus hudsonicus Meadow jumping mouse Giardia

Napaeozapus insignis Woodland jumping mouse Giardia

Peromyscus leucopus White-footed mouse Cryptosporidium, Giardia

P. maniculatus Deer mouse Cryptosporidium, Giardia

C. gapperi Red-backed vole Cryptosporidium, Giardia

Pitymys savii Savi’s woodland vole Giardia

R. norvegicus Norway rat Cryptosporidium, Giardia

Neotoma cinerea Wood rat Giardia

Dipodomys heermanni Kangaroo rat Giardia

Tamia striatus Eastern chipmunk Cryptosporidium, Giardia

Sciurus carolinensis Eastern gray squirrel Cryptosporidium, Giardia

Tamiasciurus hudsonicus Red squirrel Cryptosporidium, Giardia

Glaucomys volans Southern flying squirrel Giardia

continued

Appendix D
222

Protozoan Pathogens Giardia and Cryptosporidium

Appendix table D.1—Species reported as hosts to protozoans Giardia and Cryptosporidium (continued)

Species Common name Parasites hosted

Mammelia (continued)

 Rodentia (continued)

Spermophilus beecheyi Ground squirrel Giardia

S. richardsoni Richardson’s ground squirrel Giardia

S. tridecemlineatus 13-lined ground squirrel Giardia

Marmota monax Woodchuck Cryptosporidium, Giardia

Coendu villosus Tree porcupine Giardia

Carnivora

Ursus americanus Black bear Cryptosporidium

Mustela erminea Short-tailed weasel Cryptosporidium

M. putorius furo Ferret Cryptosporidium

Canis latrans Coyote Cryptosporidium, Giardia

Vulpes vulpes Red fox Giardia

Urocyon cineroargenteus Gray fox Cryptosporidium, Giardia

Procyon lotor Raccoon Cryptosporidium, Giardia

Paradoxurus h. hermaphroditus Palm civet Giardia

Mephitis mephitis Striped skunk Cryptosporidium, Giardia

Mustela vision Mink Cryptosporidium, Giardia

M. nigripes Black-footed ferret Giardia

Meles meles Badger Giardia

Lynx rufus Bobcat Cryptosporidium, Giardia

Sirenia

Dugong dugong Manatee Cryptosporidium

Ruminants

Cervus canadensis Elk, wapiti Giardia

Odocoileus virginiana White-tailed deer Cryptosporidium, Giardia

Antilocapra americana Pronghorn Giardia

Ovis canadensis x. O. musimon Bighorn x mouflon sheep Giardia

Llama glama Llama Cryptosporidium, Giardia

Odocoileus hemionus Mule deer Cryptosporidium

Source: Adapted from Wade and others, in press.
223

significance began to be recognized in the 1970’s when a

number of reports identified Cryptosporidium as the cause

of diarrhea in calves (O’Donoghue 1995). Human infections

began to be reported in the mid-1970’s and by the 1990’s,

Cryptosporidium was recognized as a significant threat to

individuals that are immunocompromised (Current and

Garcia 1991, Ungar 1990).

Cryptosporidium oocysts are spherical and 4 to 6 microns in

diameter (Barer and Wright 1990, Casemore 1991,

Casemore and others 1985, Current 1987, Current and

others 1986, Fayer and Ungar 1986, Issac-Renton and others

1987, O’Donoghue 1995, Smith and Rose 1990, Ungar

1994). Most oocysts contain up to four sporozoites (freeliving form). A number of species have been identified

among various hosts. Many of these species can cross-infect

different species of hosts. Several Cryptosporidium species

are found more often in association with certain host

species, especially when the host species are vertebrates.

Thus, C. muris is common in mammals, C. meleagridis in

birds, C. crotalia in reptiles, and C. nasorum in fish (Levine

1984, O’Donoghue 1995).

O’Donoghue (1995) and Current and Bick (1989) described

the life cycle for Cryptosporidium. Like Giardia,

Cryptosporidium encysts to survive outside its host, and its

life stages occur in the infected animal. Its life cycle is more

complex due to the addition of a sexual stage of reproduction within the host. The oocyst of Crypto-sporidium

undergoes excystation (release of sporozoites) after it has

been ingested by a host and has been subjected to conditions

usually found in a digestive system. These conditions have

been identified as including temperature, low pH, and

digestive enzymes (Fayer and Leek 1984, Reduker and

Speer 1985). The released sporozoites attach to epithelial

cells of the small intestine, where they are identified as

trophozoites (Cryptosporidium attached to intestine). The

trophozoites mature into meronts that produce merozoites

through asexual reproduction. The merozoites, in turn,

develop into either other meronts or produce the sexual form

of Cryptosporidium, microgametes (male form) and

macrogametes (female form). The mobile microgametes

fertilize the macrogametes in sexual reproduction to form a

zygote (the sexually reproduced form of Cryptosporidium).

Most of the zygotes form thick-walled oocysts that are

released from the host to infect other hosts and complete the

life cycle.

Literature Cited

Barer, R.R.; Wright, A.E. 1990. Cryptosporidium and water. Applied

Microbiology. 11(6): 271–277.

Campbell, I.; Tzipori, S.; Hutcinson, G.; Angus, K.W. 1982. Effects of

disinfection on survival of Cryptosporidium oocysts. Veterinary Record.

111(18): 414–415.

Casemore, D.P. 1991. Epidemiology of human Cryptosporidiosis and the

water route of infection. Water Science and Technology. 24(2): 157–164.

Casemore, D.P.; Sands, R.L.; Curry, A. 1985. Cryptosporidium species a

“ new “ human pathogen. Journal of Clinical Pathology. 38(12):

1321–1336.

Clark, R.M.; Read, E.J.; Hoff, J.C. 1989. Analysis of inactivation of

Giardia lamblia by chlorine. Journal of Environmental Engineering.

115(1): 80–90.

Clark, R.M.; Regli, S. 1993. Development of Giardia Ct values for the

surface water treatment rule. Journal of Environmental Science and

Health. 28(5): 1081–1097.

Craun, G.F. 1981. Disease outbreaks caused by drinking water. Journal of

the Water Pollution Control Federation. 53(6): 1134–1138.

Craun, G.F. 1984. Waterborne outbreaks of giardiasis. Current status. In:

Erlandsen, S.L.; Meyer, E.A., eds. Giardia and giardiasis. New York:

Plenum Press: 243–262.

Current, W.L. 1987. Cryptosporidium: its biology and potential for

environmental transmission. CRC Critical Reviews in Environmental

Control. 17(1): 21–51.

Current, W.L.; Bick, P. 1989. Immunobiology of Cryptosporidium spp.

Pathology and Immunopathology Research. 8: 141–160.

Current, W.L.; Garcia, L.S. 1991. Cryptosporidiosis. Clinical Microbiology Reviews. 4: 325–358.

Current, W.L.; Upton, S.J.; Haynes, T.B. 1986. The life cycle of

Cryptosporidium baileyi n.sp. (Apicomplexa, Cryptosporidiidae )

infecting chickens. Journal of Protozoology. 33: 29–296.

Davis, C.; Ritchie, L.S. 1948. Clinical manifestations and treatment of

epidemic amebiasis occurring in occupants of the Mantetsu apartment

building. Tokyo: American Journal of Tropical Medicine. 28: 817–823.

Fayer, R.; Leek, R.G. 1984. The effects of reducing conditions, medium,

pH, temperature, and time on in vitro excystation of Cryptosporidium.

Journal of Protozoa. 31: 567–569.

Fayer, R.; Ungar, B.L.P. 1986. Cryptosporidium spp. and Cryptosporidiosis. Microbiol.Reviews. 50(4): 458–483.

Haas, C.N.; Heller, B. 1990. Kinetics of inactivation of Giardia lamblia by

free chlorine. Water Research. 24(4): 233–238.

Hoff, J.C.; Rubin, A.J. 1987. Use of inactivation rate data for establishing

drinking water disinfection requirements for Giardia cyst control. In:

Preprints of papers presented at the 194th American chemical society

national meeting; [Date of meeting unknown]; Washington, DC.

Washington, DC: American Chemical Society, Division of Environmental Chemistry: 760–763. Vol. 27, no.2.

Appendix D
224

Reduker, D.W.; Speer, C.A. 1985. Factors influencing excystation of

Cryptosporidium oocysts from cattle. Journal of Parasitology. 71(1):

112–115.

Rice, G.R. 1981. Use of ozone in drinking water treatment. Journal

American Water Works Association. 73(1): 44.

Rose, J.B.; Gerba, C.P. 1991. Use of risk assessment for development of

microbial standards. Water Science and Technology. 24(2): 29–34.

Rose, J.B.; Haas, C.N.; Regli, S. 1991. Risk assessment and control of

waterborne Giardiasis. American Journal of Public Health. 81(6):

709–713.

Rubin, A.J.; Evers, D.P.; Eyman, C.M.; Jarroll, E.L. 1989. Inactivation

of gerbil-cultured Giardia lamblia cysts by free chlorine: Applied and

Environmental Microbiology. 55(10): 2592–2594.

Smith, H.V.; Rose, J.B. 1990. Waterborne Cryptosporidiosis. Parasitology

Today. 6(1): 8–12.

Tyzzer, E.E. 1910. An extracellular coccidium, Cryptosporidium muris, of

the gastric glands of the common mouse. Journal of Medical Research.

23: 487–509.

Ungar, B.L.P. 1990. Cryptosporidiosis in humans (Homo sapiens). Boston:

CRC Press: 59–82.

Ungar, B.L.P. 1994. Cryptosporidium and Cryptosporidiosis. Textbook of

aids medicine. Baltimore: Williams and Wilkins: 323–343.

Wade, S.E.; Ziegler, P.E.; Schaff, S.L. [In press]. Preliminary investigation of Giardia species and Cryptosporidium species in wildlife in the

New York City watershed. Final report. Ithaca, NY: Cornell University,

College of Veterinary Medicine, Department of Population Medicine and

Diagnostic Science.

Issac-Renton, J.L.; Fogel, D.; Stibbs, H.H.; Ongerth, J.E. 1987. Giardia

and Cryptosporidium in drinking water. The Lancet. April: 973–974.

Jarroll, E.L.; Bingham, A.K.; Meyer, E.A. 1981. Effect of chlorine on

Giardia lamblia cyst viability. Applied and Environmental Microbiology. 41(2): 483–487.

Jokipii, L.; Pohjola, S.; Jokipii, A.M.M. 1983. Cryptosporidium: a

frequent finding in patients with gastrointestinal symptoms. The Lancet.

2(8346): 358–842.

Kenney, R.T. 1994. Parasitic causes of diarrhea. Pediatric Annals. 23(8):

414–421.

Kong, L.I.; Swango, L.J.; Blagburn, B.L. [and others]. 1988. Inactivation of Giardia lamblia and Giardia canis cysts by combined and free

chlorine. Applied and Environmental Microbiology. 54(10): 2580–2582.

Levine, N.D. 1984. Taxonology and review of the Cooidian Genus

Cyrptosporidium (Protozoa, Aploomiexa). Journal of Protozoology. 31:

94–98.

Medema, G.J.; Van Asperen, I.A.; Klokman Houweling, J.M. [and

others]. 1995. The relationship between health effects in triathletes and

microbiological quality of freshwater. Water-Science-Technology.

31(5–6): 19–26.

Meyer, E.A.; Jarroll, E.L. 1980. Giardia and giardiasis. American Journal

of Epidemiology. 111: 1–12.

Moore, A.C.; Herwaldt, B.L.; Craun, G.F. [and others]. 1994.

Waterborne disease in the United States, 1991 and 1992. Journal

American Water Works Association. February: 87–99.

O’Donoghue, P.J. 1995. Cryptosporidium and Cryptosporidiosis in man

and animals. International Journal for Parasitology. 25(2): 139–195.

Quinn, C.M.; Betts, W.B. 1993. Longer term viability status of chlorinetreated Crypyosporidium oocysts in tap water. Biomedical Letter. 48:

315–318.

Protozoan Pathogens Giardia and Cryptosporidium
225

Appendix E

Water Treatment Technologies Tables

Gary Logsdon1

Most raw water is not suitable for human consumption

without treatment. Some water only needs to be filtered and

disinfected before consumption (Committee on Small Water

Supply Systems, National Research Council 1997). Other

water must be treated with additional processes to remove

specific chemical contaminants or nuisance chemicals like

iron and manganese. Appendix tables E.1 to E.4 present

information on water treatment techniques that can be used

for controlling common contaminants. The tables provide

guidance on selecting the appropriate treatment processes.

However, a water treatment specialist must select the best

process on a site-specific basis. Additional information can

be found in recent volumes of Water Quality and Treatment

(Letterman 1999) and Safe Water from Every Tap (Committee on Small Water Supply Systems, National Research

Council 1997).

1

 Director, Black and Veatch, Water Process Research, Cincinnati, OH.

Literature Cited

Committee on Small Water Supply Systems, National Research

Council. 1997. Safe water from every tap: improving water service to

small communities. Washington, DC: National Academy Press. 230 p.

Letterman, R.D., tech. ed. 1999. Water quality and treatment. 5th ed.

New York: McGraw-Hill, Inc. 1,248 p.
226

Water Treatment Technologies Tables

Appendix table E.1—Water treatment technologies by disinfectants, oxidants, and aerationa

General water Free Chlorine Potassium Ultraviolet

quality-constituent chlorine Chloramine dioxide Ozone permanganate radiation Aeration

General water qualities

Turbidity, sediment

Color x x x x

Disinfection by-product

precursors

Taste and odor x x x x x

Biological contaminants

Algae

Protozoa x x x x

Bacteria x x x x x

Viruses x x x x x

Organic chemicals

Volatile organics x

Semi-volatile compounds

Pesticides and herbicides

Biodegradable organic

matter

Inorganic chemicals

Hardness

Ironb x xx x x

Manganeseb x xx x

Arsenic

Selenium

Thallium

Fluoride

Radon x

Radium

Uranium

Cations

Anions

Total dissolved solids

Nitrate

Ammonia

a The columns and rows lacking x’s are where process is not appropriate or recommended for the constituent.

b When oxidant is followed by filtration.

Source: Table adapted from Committee on Small Water Supply Systems, National Research Council 1997.
227

Appendix table E.2—Water treatment technologies by type of adsorption and ion exchange systema

General water Powdered Granulated Ion Activated

quality-constituent activated carbon activated carbon exchange alumina

General water qualities

Turbidity, sediment

Color x x

Disinfection by-product

precursors x x

Taste and odor x x

Biological contaminants

Algae x

Protozoa x

Bacteria x

Viruses x

Organic chemicals

Volatile organics x x

Semi-volatile compounds x x

Pesticides and herbicides x x

Biodegradable organic

matter x x

Inorganic chemicals

Hardness x

Iron

Manganese

Arsenic x

Selenium x

Thallium x

Fluoride x x

Radon

Radium x

Uranium x

Cations x

Anions x

Total dissolved solids

Nitrate x

Ammonia

a The columns and rows lacking x’s are where process is not appropriate or recommended for the constituent.

Source: Table adapted from Committee on Small Water Supply Systems, National Research Council 1997.

Appendix E
228

Water Treatment Technologies Tables

Appendix table E.3—Water treatment technologies by type of membrane treatment systema

General water Reverse Electrodialysis/

quality-constituent Microfiltration Ultrafiltration Nanofiltration osmosis ED reversal

General water qualities

Turbidity, sediment x x x

Color x x x

Disinfection by-product

precursors x x x

Taste and odor

Biological contaminants

Algae x x x

Protozoa x x x x

Bacteria x x x

Viruses x x

Organic chemicals

Volatile organics

Semi-volatile compounds x

Pesticides and herbicides x x

Biodegradable organic

matter

Inorganic chemicals

Hardness x x x

Iron x

Manganese x

Arsenic x x

Selenium x x

Thallium x x

Fluoride x x

Radon

Radium x x

Uranium x x

Cations x x

Anions x x

Total dissolved solids x x

Nitrate x x

Ammonia

ED = electrodialysis.

a The columns and rows lacking x’s are where process is not appropriate or recommended for the constituent.

Source: Table adapted from Committee on Small Water Supply Systems, National Research Council 1997.
229

Appendix table E.4—Water treatment technologies by type of filtration systema

General water Direct Conventional Dissolved Precoat Slow sand Bag/cartridge Lime

quality-constituent filtration filtration air flotation filtration filtration filters softening

General water qualities

Turbidity, sediment x x x x x x

Color x x x

Disinfection by-product

precursors x x x

Taste and odor x

Biological contaminants

Algae x x x

Protozoa x x x x x x x

Bacteria x x x x x x

Viruses x x x x x x

Organic chemicals

Volatile organics

Semi-volatile compounds

Pesticides and herbicides

Biodegradable organic

matter x* x* x* x

Inorganic chemicals

Hardness x

Iron x x x x x

Manganese x x x x x

Arsenic x x x

Selenium x

Thallium

Fluoride

Radon

Radium x

Uranium

Cations x

Anions

Total dissolved solids

Nitrate

Ammonia x* x* x* x

X* = when the filter is operated in a biologically active mode.

a The columns and rows lacking x’s are where process is not appropriate or recommended for the constituent.

Source: Table adapted from National Research Council 1997.

Appendix E
230

List of Figures

Page

Figure 2.1—Global nitrogen cycle. Annual fluxes in units of 1012 grams per year .................................................................. 18

Figure 2.2—Sources and pathways of nitrogen in the subsurface environment ....................................................................... 19

Figure 2.3—Global phosphorus cycle. Annual fluxes in units of 1012 grams per year ............................................................. 21

Figure 3.1—Hydrologic cycle within a watershed .................................................................................................................... 27

Figure 3.2—Average acidity (pH) of precipitation in the United States from 1988 through 1997

(National Atmospheric Deposition Program 1999) ................................................................................................................... 28

Figure 3.3—The acid buffering capacity (total alkalinity) in lake water and, by inference, the

surrounding water in soils and ground water for the United States. Where lakes have low or

negative alkalinities in the black areas, surface water pH values may be <4.5. (Map prepared

by J.M. Omernick, G.E. Griffith, J.T. Irish, and C.B. Johnson with the U.S. Environmental

Protection Agency) .................................................................................................................................................................... 29

Figure 3.4—Schematic illustrating ground water terms and concepts ...................................................................................... 30

Figure 4.1—Hypothetical river basin ........................................................................................................................................ 44

Figure 4.2—Efficient allocation of cost of meeting drinking water quality standard, with one

emitter and one receptor ............................................................................................................................................................ 45

Figure 4.3—Construction cost of pressure filtration plant with an infiltration rate of 2 gallons

per minute per square foot (Gummerman and others 1978)...................................................................................................... 46

Figure 4.4—Long-term marginal cost as a function of pollutant concentration ....................................................................... 47

Figure 9.1—Surfaces and flow paths associated with a road cross section .............................................................................. 85

Figure 9.2—Mean soil loss rates for four road surfaces before, during, and for 2 years after

logging (based on Swift 1984a) ................................................................................................................................................. 89

Figure 9.3—Summary of causes of pipeline accidents in 1998 in the United States

(U.S. Department of Transportation 1999) ................................................................................................................................ 96

Figure 11.1—Mean monthly concentrations (flow weighted) of nitrate (NO3

-1) in streamwater

of a clearcut, cable-logged, hardwood-covered watershed (WS7) and an adjacent watershed

(WS2) during calibration, treatment activities, and postharvest period, Coweeta Hydrologic

Laboratory, North Carolina ....................................................................................................................................................... 121

Figure 11.2—Cumulative sediment yield measured on a clearcut, cable-logged, hardwoodcovered watershed: (A) in one of the first-order streams below a logging road during the first

32 months after treatment and (B) in the ponding basin of the second-order stream at the

gaging site during 15 years after treatment. Predicted values are based on pretreatment

calibration of sediment yield with an adjacent control watershed, Coweeta Hydrologic

Laboratory, North Carolina ....................................................................................................................................................... 123

Figure 12.1—Relationship between turbidity in nephelometric turbidity units (NTU)

and suspended sediment parts per million (ppm) (Helvey and others 1985) ............................................................................ 126

Figure 12.2—Possible pathways of plant- and litter-contained nutrients in response

to combustion (Tiedemann 1981) .............................................................................................................................................. 128
231

Page

Figure 14.1—Potential sediment production in 10 Blue Mountain ecosystems in Oregon.

Ecosystems: L = larch, M = meadow, LP = lodgepole pine, DF = Douglas fir, A = alpine,

PP = ponderosa pine, SF = spruce-fir, G = grassland, SA = sagebrush, J = juniper

(Buckhouse and Gaither 1982). Different lower-case letters indicate differences in statistical

significance (P<0.10) ................................................................................................................................................................ 154

Appendix figure A.1—Watersheds supplying Baltimore, MD, with water............................................................................... 210

Appendix figure B.1—Location of Quabbin Reservoir and water supply system

for Boston, MA .......................................................................................................................................................................... 213

Appendix figure C.1—Watershed boundary and stream sampling locations in the Coweeta

Creek Watershed in western North Carolina. Stations 1 through 5 are arranged down the

stream gradient on Coweeta Creek. First-order streams are not shown .................................................................................... 215

Appendix figure C.2—Mean and standard error (bars) for turbidity, plotted against building

density for each sampling condition (baseflow and stormflow) ............................................................................................... 216

Appendix figure C.3—Mean and standard error (bars) for (A) total coliform and (B) fecal

streptococcus, plotted against building density for each sampling condition (baseflow and

stormflow). (Building density increases downstream) .............................................................................................................. 217

List of Figures
232

List of Tables

Page

Table 2.1—Summary of common water pollutants by land-use activities ............................................................................... 8

Table 2.2—Common types of water contaminant guidelines for different water uses ............................................................. 8

Table 2.3—National primary drinking water regulations (States are expected to focus attention

on risks related to the contaminants listed in their source water assessments.) ........................................................................ 9

Table 2.4—National secondary drinking water regulations, which are nonenforceable guidelines

for contaminants that may cause cosmetic effects (e.g., skin or tooth discoloration) or aesthetic

effects (e.g., taste, odor, or color) in drinking water ................................................................................................................. 14

Table 2.5—Common chemical processes involved as water interacts with its environment ................................................... 15

Table 2.6—Summary of the chemical behavior of important water contaminants................................................................... 16

Table 2.7—Common waterborne pathogenic and indicator bacteria and viruses ..................................................................... 23

Table 2.8—Factors influencing virus transport and fate in the subsurface ............................................................................... 24

Table 6.1—Reported acres inside national forest boundaries and percent population changes in

counties containing or adjacent to national forests by administrative Forest Service Region in the

conterminous United States....................................................................................................................................................... 63

Table 6.2—Estimated use from freshwater surface and ground water sources in the

United States, 1980–95 ............................................................................................................................................................. 64

Table 6.3—Waterborne disease outbreaks in the United States by water supply system, 1990–94 ......................................... 64

Table 6.4—Comparison of outbreak percentages by drinking water source from pathogenic

contamination for the period 1971–96 ...................................................................................................................................... 64

Table 9.1—Typical erosion rates observed for different types of land use in the United States .............................................. 88

Table 9.2—Effectiveness of erosion mitigation techniques ...................................................................................................... 89

Table 9.3—Management options for decommissioned roads ................................................................................................... 91

Table 9.4—Pollutants that have been observed in runoff from road surfaces .......................................................................... 93

Table 9.5—Mean concentrations of a number of pollutants in highway runoff in Minnesota ................................................. 93

Table 9.6—Extent of soil contamination by 53 oil spills of various sizes in Alberta, Canada ................................................. 96

Table 10.1—Effects of various timber harvests or site preparations on soil erosion and

sediment production .................................................................................................................................................................. 105

Table 10.2—Effects of timber harvesting with and without streamside buffers

on stream temperature ............................................................................................................................................................... 109

Table 10.3—Effects of clearcutting with and without buffers on mean annual nitrate-nitrogen,

ammonium-nitrogen, and total-phosphorus concentrations ...................................................................................................... 112

Table 10.4—Effects of forest fertilization on maximum streamwater ammonium-nitrogen and

nitrate-nitrogen concentrations.................................................................................................................................................. 114

Table 12.1—Water turbidity, in Jackson turbidity units (equivalent to nephelometric turbidity

units), after fire alone or in combination with other treatments ................................................................................................ 125

Table 12.2—Suspended sediment concentration in streamflow after fire alone or in

combination with other treatments ............................................................................................................................................ 125
233

Page

Table 12.3—Sediment yield after fire alone or in combination with other treatments ............................................................. 127

Table 12.4—The pH in water after fire alone or in combination with other treatments ........................................................... 130

Table 12.5—Maximum nitrate-nitrogen concentration in water after fire alone or in

combination with other treatments ............................................................................................................................................ 131

Table 12.6—Sulfate concentration in water after fire alone ..................................................................................................... 133

Table 12.7—Chloride concentration in water after fire alone ................................................................................................... 133

Table 13.1—Management uses of pesticides commonly used on national forests ................................................................... 141

Table 13.2—Estimates of safe levels for daily exposure to the 20 pesticides most used

on National Forest System lands in fiscal year 1997 in the vegetation management program ................................................ 142

Table 13.3—Frequency and occurrence of surface and ground water contamination

from pesticide use in North America ........................................................................................................................................ 144

Table 15.1—Streamwater fecal coliform and Giardia responses to herbivore use................................................................... 161

Table 16.1—Indicator and principal pathogens of concern in contaminated drinking water ................................................... 165

Appendix table C.1—Summary data for the catchments above five sampling stations

along Coweeta Creek in western North Carolina ...................................................................................................................... 216

Appendix table C.2—Summary water-quality data from baseflow grab samples (means) at

each of the five sampling stations along Coweeta Creek in western North Carolina ............................................................... 217

Appendix table C.3—Summary water-quality data from stormflow samples (means) at each

of the five sampling stations along Coweeta Creek in western North Carolina ....................................................................... 217

Appendix table D.1—Species reported as hosts to protozoans Giardia and Cryptosporidium ................................................ 219

Appendix table E.1—Water treatment technologies by disinfectants, oxidants, and aeration.................................................. 226

Appendix table E.2—Water treatment technologies by type of adsorption and

ion exchange system ................................................................................................................................................................. 227

Appendix table E.3—Water treatment technologies by type of membrane treatment system .................................................. 228

Appendix table E.4—Water treatment technologies by type of filtration system ..................................................................... 229

List of Tables
234

Glossary of Abbreviations and Acronyms

ai: active ingredient

aum: animal unit month

BAER: Burn Area Emergency Rehabilitation

BLM: Bureau of Land Management

BMP: best management practice

BOD: biological oxygen demand

Bt: Bacillus thuringiensis

BTEX: benzene, toluene, ethylbenzene, xylene

C: carbon

Ca: calcium

Ca2+: calcium ion

Ca(HCO3)

2

: calcium bicarbonate

Cd: cadmium

CDC: Centers for Disease Control and Prevention

CERLA: Comprehensive Environmental Response,

Compensation and Liability Act

CH4

: methane

Cl: chlorine

Cl-1: chloride ion

cm: centimeter

CO2: carbon dioxide

COD: chemical oxygen demand

Cu: copper

CWD: coarse woody debris

DEP: City of New York Department of Environmental

Protection

DO: dissolved oxygen

DOD: Department of Defense

EPA: U.S. Environmental Protection Agency

ET: evapotranspiration

F: fluorine

FC: fecal coliform

FDA: U.S. Food and Drug Administration

Fe+2: ferrous iron ion

Fe+3: ferric iron ion

Fe3S4: greigite

FeS: pyrite

FeS2: marcasite

FIFRA: Federal Insecticide, Fungicide,

and Rodenticide Act

FS: fecal streptococcus

GIS: Geographic Information System

H+: hydrogen ion

H2S: hydrogen sulfide

ha: hectare

HA or HAL: health advisory level

HCN: hydrogen cyanide

HCO3

-1: bicarbonate ion

HTH: chlorine

JTU: Jackson turbidity unit

K: potassium
235

K+: potassium ion

km: kilometer

LOAEL: lowest observed adverse effect level

m: meter

MCL: maximum contaminant level

MCLG: maximum contaminant level goal

MD-DNR: Maryland Department of Natural Resources

MDC-DWM: Metropolitan District Commission’s

Division of Watershed Management (Massachusetts)

Mg: magnesium

Mg2+: magnesium ion

mg: milligram

Mg: metric tonne or megagram

μg: microgram

μm: micron or 1 millionth of a meter

Mg(HCO3)2: magnesium bicarbonate

MgSO4: magnesium sulfate

MITC: methyl isothiocyanate

Mn: manganese

MTBE: methyl tertiary butyl ester

N: nitrogen

N2: nitrogen gas

N2

O: nitrous oxide

Na: sodium

Na+: sodium ion

NAWQA: National Water-Quality Assessment Program

NEPA: National Environmental Policy Act

NFS: National Forest System

NH4

+1: ammonium ion

NH4+-N: ammonium-nitrogen

NO2

-1: nitrite ion

NO3

-1: nitrate ion

NO3-N: nitrate-nitrogen

NOAEL: no observed adverse effect level

NOEL: no observed effect level

NTU: nephalometric turbidity unit

O: oxygen

O2: oxygen gas

OPS: Office of Pipeline Safety

OSM: Office of Surface Mining Reclamation and

Enforcement

P: phosphorus

PAH: polycyclic aromatic hydrocarbon

Pb: lead

PFC: Proper Functioning Condition

pH: the negative logarithm (base 10) of the hydrogen

ion concentration

PO4

-3: phosphate ion

PO4

3—P: phosphate phosphorus

ppb: parts per billion

ppm: parts per million

Glossary of Abbreviations and Acronyms
236

RCRA: Resource Conservation and Recovery Act

RfD: reference dose

S: sulfur

SDWA: Safe Drinking Water Act Amendments of 1996

Si: silicon

SiO2: silica

SMRCA: Surface Mining and Control Reclamation Act

SO4

-2: sulfate ion

SWA: source water assessment

SWAP: source water assessment program

SWTR: Surface Water Treatment Rule of 1989

TDS: total dissolved solids

TFM: 3-trifluoromethyl-4-nitrophenol

THM: trihalomethane

TT: treatment technique

USGS: U.S. Geological Survey

UST: underground storage tanks

VOC: volatile organic compound

WATSED: Region 1 water and sediment model

WEPP: Water Erosion Prediction Project

WPP: Watershed Protection Program

Zn: zinc

Glossary of Abbreviations and Acronyms
237

Glossary of Terms

action level: the level of contamination which, if exceeded,

triggers treatment or other requirements that a water system

must follow.

acute health effect: an immediate, i.e., within hours or

days, effect that may result from exposure to certain

drinking water contaminants.

allelopathy: a chemical defense mechanism in certain

plants to keep other plants from growing under or around

their canopy.

anions: negatively charged ions. The most common in

natural waters are bicarbonate, nitrate, sulfate, chloride, and

different forms of phosphorus.

anoxic: lacking oxygen.

aquifer: a saturated, permeable geologic unit that can

transmit significant quantities of water under ordinary

conditions.

aquitard: a geologic unit that cannot transmit significant

quantities of water under ordinary conditions.

artesian well: a deep well in which water rises under

pressure from a permeable strata.

bed load: bed load is sediment too heavy to be continuously

suspended in flowing water. This material is rolled or

bounded along the stream bottom. The size of particles

making up the bed load varies with streamflow, velocity,

particle density and shape, and many other factors.

biological oxygen demand (BOD): dissolved oxygen

required to decompose biodegradable organic material in

parts per million (ppm) or milligrams per liter (mg/L).

capillary fringe: the zone between the water table and the

vadose zone where water is held within pores by capillary

forces.

cations: positively charged ions. The most common in

natural waters are calcium, sodium, potassium, and ammonium.

centralized wastewater treatment system: water treatment

system that collects wastewater and transports wastewater

via sewers to a central treatment facility.

chemical oxygen demand (COD): dissolved oxygen

required to decompose biodegradable and nondegradable

organic material in parts per million (ppm) or milligrams per

liter (mg/L).

closed well: a well that has been permanently disconnected

and capped or filled so that contamination cannot move

from the surface into the aquifer.

coliform: a group of related bacteria whose presence in

drinking water may indicate contamination by diseasecausing microorganisms.

coliphages: viruses (bacteriophage) that infect and replicate

in the bacterium E. coli and appear to be present wherever

E. coli are found. Some strains are more resistant to chlorine

disinfection than total coliforms.

commercial use: includes water for motels, hotels, restaurants, office buildings, golf courses, civilian and military

institutions, and in some areas fish hatcheries. The consumptive use of water for commercial purposes in the

United States in 1995 was estimated at 14 percent of

withdrawals and deliveries.

community water system: a water system with 15 or more

service connections and which supplies drinking water to 25

or more of the same people year-round in their residences.

compliance: the act of meeting all Federal and State

drinking water regulations.

confined aquifer: an aquifer that is between two impermeable geologic units.

consumptive use: the part of water withdrawn that is

evaporated, transpired, or incorporated into products or

crops. In many instances, the consumptive use is the

difference between the amount delivered and the amount

released.

contaminant: any substance found in water (including

microorganisms, minerals, chemicals, radionuclides, etc.)

which may be harmful to human health.

conveyance loss: the quantity of water that is lost in transit

from its source to point of use or point of return.
238

Glossary of Terms

Cryptosporidium spp.: a microorganism commonly found

in lakes and rivers which is highly resistant to disinfection.

Cryptosporidium has caused several large outbreaks of

gastrointestinal illness, with symptoms that include diarrhea,

nausea, and/or stomach cramps.

cutslope: excavated slope uphill from a road located on the

side of a steep hill.

decentralized water treatment system: onsite water

treatment facility.

disinfectant: a chemical (commonly chlorine, chloramine,

or ozone) or physical process, e.g., ultraviolet light, that kills

microoganisms such as bacteria, viruses, and protozoa.

distribution system: a network of pipes leading from a

treatment plant to customer’s plumbing system.

domestic use: includes water used for normal household

purposes, such as drinking, food preparation, bathing,

washing clothes and dishes, flushing toilets, and watering

lawns and gardens. The consumptive use of water for

domestic purposes in the United States in 1995 was estimated at 26 percent of withdrawals and deliveries.

drywell: a well used for disposal of liquid wastes, other

than an improved sinkhole or subsurface fluid distribution

system, completed above the water table so that its bottom

and sides are typically dry except when receiving fluids.

enteric viruses: viruses which infect the gastrointestinal

tract of humans and are excreted with the feces of the

infected individual. These viruses are excreted in relatively

large numbers from infected individuals and include

polioviruses, coxsackieviruses, echoviruses, other enteroviruses, adenoviruses, rotaviruses, hepatitis A virus, Norwalk

viruses, astrovirus, and caliciviruses.

equivalents per liter: a chemical term indicating the

number of moles of solute multiplied by the valence of the

solute species in 1 liter of solution.

eutrophication: enrichment of surface waters with nutrients, especially phosphorus and nitrogen that leads to

enhanced plant growth, algal blooms and depleted oxygen

levels as this plant material decays.

exotic: with reference to vegetation, refers to nonnative

plant species introduced either accidentally, or to meet some

management goal.

fillslope: the downhill embankment on a road constructed

on the side of a steep hill.

finished water: drinking water that has been treated and is

ready to be delivered to customers.

fire intensity: in a wildfire or prescribed burn, a qualitative

term describing the rate of heat release, related to flame

length.

fire severity: in a wildfire or prescribed burn, a qualitative

term describing the extent of fire effects on ecosystem

components, such as vegetation or soils.

gaining stream: a stream that receives flow from ground

water discharge.

Giardia spp.: a microorganism frequently found in rivers

and lakes, which, if not treated properly, may cause diarrhea, fatigue, and cramps after ingestion.

ground water: the water that drinking water systems pump

and treat from aquifers.

industrial use: includes water use for processing, washing,

and cooling in facilities that manufacture products like steel,

chemicals, paper, and petroleum refining. The consumptive

use of water for industrial purposes in the United States in

1995 was estimated at 15 percent of total withdrawals and

deliveries.

in holdings: land parcels contained within public lands that

are not owned by the agency managing public lands.

inorganic contaminants: mineral-based compounds, such

as metals, nitrates, and asbestos. These contaminants are

naturally occurring in some water, but can also get into

water through farming, chemical manufacturing, and other

human activities.

instream use: water use that takes place without the water

body being diverted or withdrawn from surface or ground

water sources. Examples include hydroelectric power

generation, navigation, freshwater dilution of saline estuaries, maintenance of minimum streamflows to support fish

and wildlife habitat, and wastewater assimilation.
239

irrigation use: includes all water artificially applied to farm

and horticultural crops and in some cases golf courses. Of

the water withdrawn for irrigation in the United States in

1995, 19 percent was lost in conveyance, 61 percent was

consumptive use, and 20 percent was returned to surface of

ground water supplies.

karst: geologic formation in limestone strata containing

numerous dissolved, undergound channels resulting in high

hydraulic conductivity and high risk of ground water

pollution.

leachate: a liquid, often containing extremely high concentrations of organic and inorganic pollutants, formed from the

decomposition of municipal solid waste.

livestock use: includes offstream use of water for livestock,

feed lots, dairies, fish farms, and other on-farm needs. The

consumptive use of water for livestock in the United States

in 1995 was estimated at 26 percent of withdrawals and

deliveries.

losing stream: a stream that loses water to the ground

water. Streams that help recharge ground water.

mass concentration: the mass of a solute dissolved in a

specific unit volume of solution, usually expressed in

milligrams per liter.

mass failure: the collapse of a steep embankment when the

gravitational forces within the embankment exceed the

strength of the soil to maintain the current slope.

maximum contaminant level (MCL): the highest level of a

contaminant that U.S. Environmental Protection Agency

allows in drinking water to ensure that drinking water does

not pose either a short-term or long-term health risk.

maximum contaminant level goal (MCLG): the level of a

contaminant at which there would be no risk to human

health. This goal is not always economically or technologically feasible, and the goal is not legally enforceable.

microorganisms: tiny living organisms that can be seen

only with the aid of a microscope. Some microorganisms

can cause acute health problems when consumed in drinking

water. Also known as microbes.

mining use: offstream water uses for the extraction and

milling of naturally occurring minerals including coals and

ores, petroleum, and natural gases. The consumptive use of

water for mining purposes in the United States in 1995 was

estimated at 27 percent of withdrawals and deliveries.

molality: the number of moles of solute in a 1 kilogram

mass of solvent.

molarity: the number of moles of solute in a liter of

solution.

monitoring: testing that water systems must perform to

detect and measure contaminants. A public water system

that does not follow U.S. Environmental Protection

Agency’s monitoring methodology or schedule is in

violation, and may be subject to legal action.

municipal solid waste landfill: a discrete area of land or an

excavation that receives household wastes.

nonpoint-source pollution: contaminants that come from

diffuse sources and pollute surface and ground water

sources.

nontransient, noncommunity water system: a system

which supplies drinking water to 25 or more of the same

people at least 6 months per year in places other than their

residences. Some examples are schools, factories, office

buildings, and hospitals that have their own water systems.

offstream use: water that is diverted or withdrawn from

surface or ground water sources and conveyed to the place

of use.

organic contaminants: carbon-based chemicals, such as

solvents and pesticides, which can get into water through

runoff from cropland or discharge from factories. U.S.

Environmental Protection Agency has set legal limits on

50 organic contaminants.

oxygenates: organic compounds added to gasoline to

increase oxygen content of gasoline and reduce certain

emissions.

pathogen: a disease-causing organism.

perched water table: a zone of saturation that is bound

below by impermeable material elevated above a vadose

zone above the water table.

Glossary of Terms
240

Glossary of Terms

pH: a common measure of acidity and alkalinity defined as

the negative logarithm (base 10) of the hydrogen ion

concentration. A pH of 7 represents neutral conditions, a pH

value <5 indicates moderately acidic conditions, while a pH

value >9 indicates moderately alkaline conditions.

plugging: the act or process of stopping the flow of water,

oil, or gas into or out of a formation through a borehole or

well penetrating that formation.

point-source pollution: contaminants that can be traced to

specific points of discharge and pollute surface and ground

water sources.

polycyclic aromatic hydrocarbons (PAH): multiple-ringed

carbon compounds that are potentially carcinogenic.

potentimetric surface: the water surface level of the

saturated zone in a confined aquifer.

public supply: water withdrawn by public and private

suppliers and delivered to multiple users for domestic,

commercial, industrial, and thermoelectric power uses. The

difference between the amount of water withdrawn and

delivered to users typically represents losses in the distribution system and use for water treatment plant filter cleaning,

water for fire fighting, street cleaning, and occasionally

municipal buildings.

public water system: any water system which provides

water to at least 25 people for at least 60 days.

radionuclides: any man-made or natural element that emits

radiation and that may cause cancer after many years of

exposure through drinking water.

raw water: water in its natural state, prior to any treatment

for drinking.

redox potential: a measure of the oxidizing or reducing

capacity of a solution where positive values indicate

oxidizing tendencies and negative values indicate reducing

tendencies. Chemically, it is defined as the energy gained in

the transfer of 1 mole of electrons from an oxidant to

hydrogen.

return flow: the quantity of water that is discharged to a

surface or ground water after release from the point of use

and, thus, becomes available for further uses.

road prism: the road and surrounding area directly influenced by the road, including any cutslopes, ditches,

fullslopes, and the roadway.

roadway: the surface of a road on which vehicles travel.

rock buttress: a thick layer of rock placed on top of a steep

sideslope to reduce the risk of a mass failure.

saturated zone: a soil or geologic zone in which all pores

are filled with water.

secondary drinking water standards: nonenforceable

Federal guidelines regarding cosmetic effects (e.g., tooth or

skin discoloration) or aesthetic effects (e.g., taste, odor, or

color) of drinking water.

slumping: a mass failure generally due to an increase in

water content within the soil profile on steep slopes.

sole-source aquifer: an aquifer that supplies 50 percent or

more of the drinking water of an area.

spring: places on the land surface where the water table or

an aquifer intersects the land surface, discharging ground

water.

surface water: the water in sources open to the atmosphere,

such as rivers, lakes, and reservoirs.

suspended sediment: suspended sediment is material light

enough to be carried in suspension in streamflow. The

sediment carried in suspension may be either organic or

inorganic material. Unless specified, both types are included

in suspended sediment estimates. Suspended sediment is

often reported as the concentration in water using parts per

million or milligrams per liter interchangeably to express the

instantaneous concentration at a given point. Sediment not

transported in suspension is called bed load.

thermoelectric power: includes offstream uses for the

generation of electric power with fossil fuels, nuclear, or

geothermal power. In the United States in 1995, surface

water supplied more than 99 percent of the thermoelectric

withdrawals. Consumptive use was about 2 percent of

withdrawals.
241

total dissolved soils (TDS): determined by weighting the

solid residue obtained by evaporating a measured volume of

filtered water. Reported in mass per-unit volume, typically

in milligrams per liter.

total sediment yield: total sediment yield includes both

suspended sediment yield and bed load yield at a point along

a stream.

transient, noncommunity water system: a water system

which provides water in a place such as a gas station or

campground where people do not remain for long periods of

time. These systems do not have to test or treat their water

for contaminants that pose long-term health risks because

fewer than 25 people drink the water over a long period.

They still must test their water for microbes and several

chemicals that pose short-term risks.

turbidity: turbidity of water is the degree to which light

penetration is impeded by suspended material. Turbidity is

expressed either in Jackson turbidity units (JTU) or nephelometric turbidity units (NTU).

turbidity unit (tu): one tu is the interference in the passage

of light caused by a suspension of 1 milligram per liter of

silica. Turbidity ≤5 tu is generally not noticeable to the

average person.

unconfined aquifer: or water table aquifer. An aquifer in

which the water table forms the upper boundary.

vadose zone: this is a geologic or soil zone, which is not

saturated. It is a zone of aeration, and water in this zone

follows the laws of soil physics. Water in this zone does not

flow to a well.

volatile organic compounds (VOC): organic compounds

that volatilize at room temperatures.

vulnerability assessment: an evaluation of drinking water

source quality and its vulnerability to contamination by

pathogens and toxic chemicals.

wastewater release: includes the disposal of water conveyed through a sewer system. In the United States in 1995,

approximately 2 percent of these releases were reclaimed for

beneficial uses, such as irrigation of golf courses and public

parks.

water bar: a ditch excavated across a road to route water

from the road surface or uphill ditch to a downhill ditch or

hillside, to reduce surface erosion by concentrated flow, and

distribute surface runoff along a hillside.

water birds: water birds refer to all waterfowl including

swans, geese, and ducks (whistling ducks, marsh ducks,

diving ducks, stiff-tailed ducks, and mergansers) or ducklike swimming birds including gulls, cormorants, grebes,

loons, coots, and wading birds, such as herons and egrets.

water delivery: the quantity of water delivered to a specific

point of use.

water release: the quantity of water released to surface

water or ground water after a specific use.

water table: the level at which water stands in a well and is

the point where fluid pressure in the pores is exactly

atmospheric. Also called phreatic surface.

watershed: the land area from which water drains into a

stream, river, lake, or reservoir.

well: a bored, drilled or driven shaft, or a dug hole, whose

depth is greater than its largest surface dimension.

wellhead protection area: the area surrounding a drinking

water well or well field, which is protected to prevent

contamination of the well(s).

withdrawal: the quantity of water diverted or withdrawn

from a surface or ground water source.

Glossary of Terms
242

Subject Index

abandoned wells ......................... 62, 65, 68, 69, 192, 204

acid ............. 7, 14, 16, 17, 21, 28, 29, 31, 85, 92, 94,

111, 170, 174, 179, 180, 181, 182, 183, 186, 188,

189, 191, 204, 230

agriculture ............ 4, 8, 36, 42, 43, 46, 48, 55, 90, 147,

153, 166, 192, 215

algae bloom .................................................. 23, 56, 57, 60

algicides ........................................................................... 140

Alnus ......................................................... 39, 58, 113, 215

ammonia .................. 14, 17, 20, 58, 76, 113, 132, 169,

173, 226, 227, 228, 229

antibacterials .................................................................... 170

aquaculture .............................................................. 169, 170

aquifers ................... 24, 57, 58, 59, 65, 68, 69, 71, 77,

147, 182, 190, 191, 192, 193, 204, 205, 238

Army Corps of Engineers .................................................. 55

auto emissions ................................................................... 78

bacteria ..................... 8, 13, 14, 16, 17, 20, 22, 23, 24,

31, 34, 58, 59, 64, 65, 66, 76, 82, 96, 104, 154, 155,

158, 164, 165, 167, 169, 183, 185, 188, 216,

217, 226, 227, 228, 229 232, 237, 238

BAER ............................................................ 126, 134, 234

Baltimore ................................................................ 209, 210

beaver ..................................................................... 155, 160

benzene, toluene, ethylbenzene, and xylene ............. 75, 234

best management practice........................................... 5, 234

biological oxygen demand ... 8, 48, 55, 93, 169, 234, 237

BLM ................... 96, 153, 154, 184, 190, 191, 193, 234

BMP ............................................................................ 5, 234

BOD ..................................... 8, 48, 55, 93, 169, 234, 237

Boston ..................................................................... 212, 213

BTEX ........................................................................ 75, 234

buffers ........................ 93, 108, 109, 110, 112, 113, 114,

115, 129, 139, 146, 148, 204, 232

Bureau of Land Management .............. 96, 153, 154, 184,

190, 191, 193, 234

Bureau of Reclamation ...................................................... 55

Burn Area Emergency Rehabilitation ............ 126, 134, 234

calcium ............... 7, 20, 28, 31, 33, 34, 58, 64, 92, 111,

120, 128, 170, 215

campgrounds................................................... 3, 70, 74, 75

campylobacter ............................................................. 23, 81

cations ................ 7, 14, 17, 20, 24, 28, 31, 92, 94, 128,

226, 227, 228, 229, 237

cattle ....... 11, 12, 43, 56, 154, 155, 158, 159, 162, 204

CDC .............................................................. 6, 78, 79, 234

Centers for Disease Control and Prevention ...... 6, 78, 79,

234

CERLA ................................................................... 184, 234

chemical oxygen demand ...................................... 8, 55, 56

chemical spills ............................... 95, 199, 202, 204, 205

chloride ..................... 7, 12, 13, 14, 16, 20, 31, 58, 92,

93, 128, 132, 133, 134, 170, 215, 233, 234, 237

chlorine ........... 14, 32, 46, 57, 76, 170, 212, 226, 234,

237, 238

City of New York Department of Environmental Protection

167, 234

Clean Water Act ............. 5, 7, 42, 49, 59, 65, 184, 189,

192, 202

clearcutting .......... 37, 110, 111, 112, 120, 128, 132, 232

climate ......... 32, 33, 35, 38, 69, 81, 82, 90, 110, 126,

148, 153, 186, 197, 198

coal ............................. 9, 14, 17, 28, 183, 187, 188, 189

coarse woody debris ................................................ 173, 234

COD ..................................................................... 8, 93, 237

coliform ........... 8, 13, 16, 22, 23, 58, 59, 64, 82, 154,

155, 158, 159, 160, 161, 164, 165, 167, 215, 216,

217, 231, 233, 234, 237

color ......................... 7, 14, 18, 20, 21, 22, 56, 128, 205

Comprehensive Environmental Response, Compensation

and Liability Act .................................................. 184, 234

contaminants .......... 3, 7, 9, 10, 11, 12, 13, 14, 16, 22,

23, 31, 42, 65, 68, 69, 71, 75, 78, 79, 85, 134,

164, 179, 181, 182, 183, 184, 185, 188, 189, 192,

193, 202, 203, 205, 206, 225

corridors ................ 85, 86, 88, 90, 92, 94, 96, 98, 100,

191, 203

costs .......................... 7, 22, 42, 43, 45, 46, 47, 48, 50,

55, 56, 57, 65, 76, 83, 104, 185, 198, 204, 212, 213

Coweeta Creek........................................................ 215, 217

Cryptosporidium ........ 22, 46, 49, 76, 81, 82, 155, 158,

165, 166, 218, 219, 220, 221, 222, 223, 233, 238

CWD ....................................................................... 173, 234

cyanide .............................. 9, 16, 180, 182, 183, 184, 186

Cyclosporidium.................................................................. 64

cysts ................................................................... 81, 82, 218

dams ............ 49, 55, 56, 57, 59, 60, 173, 180, 182, 203

deer ................ 140, 154, 155, 158, 159, 160, 204, 212,

213, 214

deicing .......................................................... 42, 49, 92, 94

DEP......................................................................... 167, 234

Department of Defense ....................................... 70, 71, 234

disinfectant ............................................... 22, 76, 170, 238
243

Subject Index

dispersed recreation .................... 74, 81, 82, 83, 203, 204

dissolved oxygen ............... 8, 14, 17, 20, 21, 22, 23, 31,

48, 234

Division of Watershed Management .... 212, 213, 214, 235

DO ........................ 8, 14, 17, 20, 21, 22, 23, 31, 48, 234

DOD .................................................................. 70, 71, 234

drilling ..................... 9, 58, 69, 190, 191, 192, 193, 204

drinking water standards.................. 7, 43, 49, 58, 66, 75,

121, 132, 141, 143, 148, 202, 240

ducks ....................................................................... 164, 167

E. coli........................................ 13, 23, 76, 158, 159, 237

economics ................................................................... 43, 47

ecosystems ..... 17, 20, 26, 32, 33, 38, 39, 76, 120, 121,

122, 129, 140, 154, 166, 172, 174, 199, 231

effluent ......... 32, 55, 58, 59, 65, 64, 78, 169, 170, 181,

199

elk ....................................... 154, 155, 158, 159, 160, 204

environment ....... 7, 14, 19, 37, 38, 45, 53, 65, 70, 93,

147, 170, 181, 184, 186, 188, 197, 202, 203, 218,

230, 232

Environmental Protection Agency .............. 3, 4, 6, 13, 14,

22, 23, 24, 42, 43, 49, 55, 59, 63, 64, 65, 66, 67,

68, 70, 74, 75, 76, 77, 124, 139, 140, 141, 142,

143, 147, 164, 167, 184, 191, 192, 193, 197, 198,

202, 210, 213, 218, 234, 234

EPA .......... 3, 4, 6, 13, 14, 22, 23, 24, 42, 43, 49, 55,

59, 63, 64, 65, 66, 67, 68, 70, 74, 75, 76, 77, 124,

139, 140, 141, 142, 143, 147, 164, 167, 184, 191,

192, 193, 197, 198, 202, 210, 213, 218, 234, 234

erosion ...................... 9, 10, 13, 22, 31, 34, 35, 36, 37,

43, 47, 50, 55, 75, 76, 77, 81, 85, 86, 87, 89, 90,

91, 92, 93, 94, 95, 96, 103, 104, 107, 108, 126,

128, 134, 135, 153, 154, 156, 173, 181, 182, 186,

189, 199, 203, 205

Escherichia coli ........................ 13, 23, 76, 158, 159, 237

ET ............................................................................. 29, 234

evapotranspiration .................................................... 29, 234

FC ... 8, 16, 23, 64, 82, 154, 158, 159, 160, 161, 164,

165, 167, 215, 216, 217, 233, 234

FDA ........................................................................ 170, 234

fecal coliform.... 8, 16, 23, 64, 82, 154, 158, 159, 160,

161, 164, 165, 167, 215, 216, 217, 233, 234

fecal streptococcus..................... 158, 215, 216, 217, 234

feces ....... 76, 81, 82, 83, 155, 156, 158, 159, 164, 170,

218

Federal Insecticide, Fungicide, and Rodenticide Act139, 234

fertilization ............ 33, 34, 111, 113, 115, 129, 160, 162

FIFRA ..................................................................... 139, 234

fire.... 9, 32, 33, 34, 36, 37, 42, 47, 55, 96, 103, 107,

111, 120, 124, 125, 126, 128, 129, 132, 134, 135,

203, 204

fire severity .................................................... 134, 135, 203

fish ............ 17, 21, 28, 42, 43, 47, 49, 50, 55, 76, 82,

129, 169, 170, 171, 172, 173, 174, 175, 183, 184,

204, 205, 223

floods ............... 26, 35, 36, 95, 173, 202, 203, 204, 205

fluoride ..... 9, 14, 16, 17, 134, 212, 226, 227, 228, 229

Forest Service ... 3, 4, 5, 6, 55, 56, 62, 63, 67, 68, 69,

71, 74, 77, 79, 81, 85, 86, 90, 91, 92, 95, 113,

114, 124, 126, 132, 139, 140, 142, 146, 154, 161,

179, 182, 184, 190, 191, 197, 205, 209, 210, 232

forest succession ............................................ 104, 121, 122

forests ........ 3, 4, 5, 8, 18, 22, 26, 30, 33, 34, 39, 42,

43, 44, 104, 108, 111, 113, 120, 121, 124, 139, 140,

141, 193, 197, 198, 200, 202, 209, 210, 232, 233

FS ............................................... 158, 215, 216, 217, 234

fumigants ....................................................... 140, 147, 148

fungicides ............................................... 65, 140, 147, 148

gas ............ 10, 14, 17, 20, 27, 32, 68, 69, 70, 75, 111,

147, 170, 177, 183, 190, 191, 192, 193, 194, 204,

235, 240, 241

gasoline ......................................................... 14, 68, 75, 78

geese ..................................................... 164, 166, 167, 213

Geographic Information System........ 197, 198, 210, 215,

234

Giardia ............. 13, 22, 42, 47, 49, 64, 76, 81, 82, 155,

158, 159, 160, 161, 162, 165, 166, 218, 219, 220,

221, 222, 223, 224, 233, 238

GIS .............................................. 197, 198, 210, 215, 234

Glyphosate ...................................................... 11, 142, 146

grazing ........ 4, 8, 36, 37, 42, 153, 154, 155, 156, 160,

197, 198, 199, 204, 205, 215

ground water ......... 6, 7, 14, 17, 18, 20, 21, 23, 24, 26,

27, 28, 29, 30, 31, 32, 39, 58, 59, 60, 62, 63, 64,

65, 67, 68, 69, 70, 71, 74, 75, 76, 77, 78, 85, 86,

92, 94, 95, 96, 111, 114, 132, 135, 139, 143, 146,

147, 148, 149, 155, 180, 181, 182, 183, 184, 185,

188, 189, 192, 193, 203, 205, 210

gulls ............................................. 164, 166, 167, 213, 241

HA......................................................... 140, 142, 146, 234

HAL ....................................................... 140, 142, 146, 234

HCN ........................................................................ 183, 234

health advisory level .............................. 140, 142, 146, 234

Helicobacter pylori ..................................................... 23, 64

hepatitis A .......................................... 23, 64, 76, 165, 238

hepatitis E .................................................................... 23, 64

herbicides .................... 9, 14, 17, 65, 103, 140, 147, 203

herbivores .................................... 158, 159, 160, 161, 162

hexazinone .............................................................. 143, 147

hunting ........................................................ 8, 74, 212, 214
244

Subject Index

hydrocarbons .... 8, 16, 65, 75, 76, 86, 92, 94, 96, 166,

192, 193

hydrogen cyanide.................................................... 183, 234

hydrology ...................... 26, 60, 76, 85, 86, 95, 108, 186

hydromodification ............................................................. 55

ice ......................................20, 29, 56, 93, 108, 111, 170

impervious surfaces ............................ 63, 67, 77, 78, 203

impoundments ............................... 56, 180, 181, 182, 186

industrial .......... 4, 8, 10, 11, 12, 14, 17, 28, 32, 39, 42,

47, 65, 68, 70, 95, 96, 139, 143, 172, 179, 238, 240

inorganic pollutants ............................................ 8, 70, 239

insecticides ............................................................... 17, 140

Jackson turbidity unit .............................................. 234, 241

JTU .......................................................................... 234, 241

kudzu ............................................................................... 140

land use ...... 5, 8, 35, 36, 37, 62, 67, 81, 89, 103, 199,

202, 209, 210, 215, 216, 217, 232

landfills ............. 9, 10, 12, 42, 62, 63, 70, 71, 166, 213

landslides ................ 34, 35, 36, 42, 85, 87, 95, 96, 202,

203, 205

leachates ................................................................... 70, 173

legionella .................................................................... 13, 23

liming ............................................................................... 174

livestock ...... 8, 12, 42, 45, 56, 59, 153, 154, 155, 156,

159, 160, 162, 188, 204, 216

LOAEL ................................................. 140, 141, 142, 235

logging ............ 36, 37, 82, 89, 90, 104, 107, 120, 122,

123, 173, 202, 209, 230

lowest observed adverse effect level .... 140, 141, 142, 235

manganese ...... 14, 16, 20, 21, 56, 132, 134, 184, 188,

225, 226, 227, 228, 229, 235

marina ................................................................................ 76

Maryland Department of Natural Resources .......... 209, 210,

235

maximum contaminant level ....... 9, 10, 11, 12, 13, 129,

134, 183, 188, 235, 239

maximum contaminant level goal ........ 9, 10, 11, 12, 13,

235, 239

MCL ........... 9, 10, 11, 12, 13, 129, 134, 183, 188, 235,

239

MCLG................................... 9, 10, 11, 12, 13, 235, 239

MD-DNR ..................................................... 209, 210, 235

MDC-DWM ......................................... 212, 213, 214, 235

mercury .............. 9, 16, 28, 93, 134, 174, 180, 184, 188

metals .......... 7, 14, 17, 28, 36, 42, 57, 59, 66, 77, 78,

86, 92, 94, 132, 172, 174, 179, 180, 181, 182, 183,

184, 185, 188, 189, 192, 204

methyl tertiary butyl ester .................... 68, 75, 78, 79, 235

Metropolitan District Commission’s Division of Watershed

Management....................................... 212, 213, 214, 235

microorganisms ........ 13, 14, 22, 24, 34, 59, 64, 81, 83,

237, 239

milling..................................................................... 180, 239

mineral extraction ................................................................ 8

mines........... 10, 17, 62, 179, 180, 181, 182, 183, 184,

185, 187, 188, 204

mining .... 14, 17, 36, 42, 95, 174, 179, 180, 181, 182,

183, 184, 185, 187, 188, 189, 197, 202, 204

moose ............................................................. 158, 159, 160

MTBE .................................................. 68, 75, 78, 79, 235

National Environmental Policy Act .............. 5, 37, 39, 235

National Forest System..... 3, 6, 74, 140, 142, 143, 147,

148, 149, 233, 235

National Water-Quality Assessment Program ........ 147, 235

NAWQA ................................................................. 147, 235

NEPA ............................................................ 5, 37, 39, 235

nephalometric turbidity unit ................. 126, 230, 235, 241

New York ....................... 47, 49, 100, 164, 166, 167, 218

NFS ........ 3, 6, 74, 140, 142, 143, 147, 148, 149, 233,

235

nitrate ....... 7, 8, 9, 16, 17, 28, 32, 33, 34, 38, 39, 58,

111, 113, 114, 115, 120, 121, 123, 128, 129, 134,

135, 155, 156, 169, 180, 182, 184, 203, 204, 210,

215

nitrites ............................................................................ 8, 17

nitrogen saturation ....... 32, 33, 111, 114, 115, 120, 129,

134, 135, 199

no observed adverse effect level ........... 140, 141, 142, 235

no observed effect level ........................ 140, 141, 142, 235

NOAEL ................................................. 140, 141, 142, 235

NOEL.................................................... 140, 141, 142, 235

Norwalk virus ............................................................. 23, 64

NTU ...................................................... 126, 230, 235, 241

nutrients .... 17, 26, 30, 31, 34, 37, 39, 42, 56, 59, 65,

66, 94, 103, 110, 111, 124, 128, 134, 135, 147,

156, 166, 172, 210

odor ................................................. 7, 14, 18, 20, 21, 205

Office of Pipeline Safety .......................................... 97, 235

Office of Surface Mining Reclamation and Enforcement

 188, 235

oil ........ 68, 69, 75, 76, 95, 96, 97, 140, 180, 190, 191,

192, 193, 204, 214

oil spills ............................................................. 76, 96, 232

oocysts ............................................. 46, 81, 155, 218, 223

OPS ........................................................................... 97, 235

ore bodies..................................... 179, 180, 183, 184, 185

organic pollutants ......................................................... 8, 31

OSM ....................................................................... 188, 235
245

PAH ........................................................................ 235, 240

parasites .................................................. 65, 169, 170, 218

pathogens ......... 22, 32, 42, 46, 47, 49, 55, 58, 64, 75,

76, 77, 82, 83, 108, 139, 140, 149, 154, 155, 158,

164, 166, 198, 199, 203, 204, 205, 213, 218

pesticides ..... 14, 20, 42, 57, 59, 65, 86, 95, 139, 140,

142, 143, 146, 147, 148, 149, 166, 198, 203

pets ..................................................... 56, 82, 83, 203, 204

PFC ......................................................................... 154, 235

pH ........ 27, 28, 31, 92, 113, 128, 129, 134, 159, 162,

169, 171, 174, 182, 183, 184, 185, 215, 223

phosphate .... 14, 56, 58, 128, 129, 132, 134, 155, 156,

169, 215, 235

phosphorus ........... 7, 14, 18, 20, 30, 33, 34, 42, 46, 48,

56, 107, 110, 111, 113, 114, 115, 128, 132, 166,

169, 215

pipeline ....................................................... 57, 95, 96, 203

pipeline accidents .............................................. 95, 96, 230

piscicides ......................................................................... 171

pollutants ......... 44, 77, 92, 93, 94, 23, 28, 31, 37, 44,

45, 47, 57, 58, 65, 69, 70, 74, 75, 76, 77, 78, 86,

87, 92, 93, 94, 129, 164, 166, 182, 184, 202, 205,

209, 210, 232, 239

polycyclic aromatic hydrocarbon ........................... 235, 240

potassium ........... 7, 58, 64, 92, 111, 120, 128, 170, 171,

215

precipitation ..... 26, 27, 28, 29, 35, 36, 42, 59, 66, 77,

78, 86, 87, 103, 107, 108, 111, 113, 115, 122, 128,

132, 135, 148, 153, 154, 155, 174, 180, 183, 184,

197, 198

prescribed burn ............................................... 89, 131, 238

Proper Functioning Condition ................................ 154, 235

protozoa ............................................................. 64, 76, 154

railroads ......................................................... 85, 90, 91, 95

rangeland .......... 17, 35, 58, 59, 60, 93, 128, 129, 132,

140, 153, 55, 156, 161, 199

raw water ......................................................... 22, 225, 240

RCRA ....................................... 69, 70, 71, 192, 193, 236

recreation ......... 32, 37, 42, 43, 44, 49, 55, 67, 74, 75,

76, 77, 78, 79, 81, 82, 83, 89, 139, 164, 197, 199,

203, 204, 209, 211, 213

Region 1 water and sediment model ........................ 90, 236

research needs ............................................. 26, 57, 78, 206

reservoirs ......... 17, 21, 23, 29, 42, 43, 49, 55, 56, 57,

60, 63, 74, 75, 76, 126, 143, 146, 159, 166, 167,

171, 191, 198, 209, 210, 212, 214

Resource Conservation and Recovery Act of 1976 .......... 69,

70, 71, 192, 193, 236

riparian ..........18, 32, 34, 38, 39, 43, 56, 63, 108, 111,

128, 129, 153, 154, 155, 156, 172, 182, 198, 205,

209, 210, 215

roads ........... 36, 37, 43, 47, 65, 77, 78, 81, 85, 86, 87,

89, 90, 91, 92, 94, 95, 99, 103, 104, 122, 173,

182, 190, 191, 192, 193, 199, 203, 215

Rotavirus................................................................... 23, 165

rotenone ........................................................................... 171

Safe Drinking Water Act Amendments of 1996 ....... 3, 4, 5,

212, 213, 218, 236

salinity .......................................................................... 8, 17

Salmonella .......................................... 23, 64, 82, 158, 165

salts ........................................................ 8, 14, 24, 92, 193

SDWA .................................... 3, 4, 5, 212, 213, 218, 236

sediment production .... 34, 35, 36, 38, 39, 90, 104, 105,

106, 107, 108, 124, 154, 231, 232

sediments ......... 20, 26, 45, 56, 57, 60, 65, 66, 76, 92,

94, 107, 122, 155, 156, 159, 162, 170, 171, 182,

184, 203, 204, 217

septic systems ......... 23, 42, 44, 45, 49, 64, 65, 67, 68,

203, 205

sewage ............. 17, 22, 32, 55, 58, 59, 60, 64, 65, 70,

76, 77, 94, 95, 96, 97, 166, 169, 170, 172

sewage treatment plants ................................ 32, 55, 59, 65

sewer systems .................................................................... 49

sheep ................................................................................ 154

Shigella .............................................................. 23, 64, 165

site preparation ... 65, 103, 104, 107, 140, 141, 143, 193

skiing ............................................................ 74, 75, 77, 78

SMRCA .................................................................. 188, 236

snow making ............................................... 74, 77, 78, 197

sodium chloride ........................................................ 92, 170

soil properties .............................................................. 24, 35

solid waste landfill ............................................. 70, 71, 239

source water assessment ................................... 3, 210, 236

Source Water Assessment Program .................. 3, 210, 236

sulfur ........ 14, 17, 20, 21, 30, 34, 126, 128, 132, 187,

188, 189, 193, 236

Surface Mining and Control Reclamation Act ....... 188, 236

Surface Water Treatment Rule of 1989 ......... 167, 213, 236

SWA ................................................................... 3, 210, 236

SWAP................................................................ 3, 210, 236

swimming ..... 47, 75, 76, 77, 79, 81, 83, 164, 171, 204

SWTR ............................................................ 167, 213, 236

taste ................................. 7, 14, 20, 21, 30, 56, 128, 205

TDS.................................................................. 31, 236, 241

temperature ........ 7, 14, 17, 20, 21, 22, 27, 29, 30, 55,

58, 59, 103, 108, 110, 124, 128, 155, 159, 162,

169, 171, 172, 190, 215, 216, 223

THM ................................................................. 56, 57, 236

timber ....... 37, 42, 45, 47, 87, 103, 104, 107, 108, 110,

111, 113, 122, 140, 188, 197, 199, 203, 204, 209,

214

timber harvests.......................................105, 106, 111, 232

Subject Index
246

total dissolved solids........................................ 31, 236, 241

toxic algae bloom .............................................................. 56

toxics................................................................................ 205

trace metals ..................................... 7, 8, 14, 17, 172, 174

traffic .................. 74, 77, 78, 79, 81, 86, 87, 90, 94, 95

trails ....................................... 47, 74, 77, 81, 85, 89, 104

treatment technique ............... 9, 10, 11, 12, 13, 213, 236

trihalomethane ................................................... 56, 57, 236

TT .......................................... 9, 10, 11, 12, 13, 213, 236

turbidity .. 20, 22, 34, 35, 36, 58, 63, 75, 77, 86, 103,

104, 107, 124, 125, 126, 128, 134, 159, 171, 203,

213, 215, 216

U.S. Department of Agriculture, Forest Service..... 3, 4, 5,

6, 55, 56, 62, 63, 67, 68, 69, 71, 74, 77, 79, 81,

85, 86, 90, 91, 92, 95, 113, 114, 124, 126, 132,

139, 140, 142, 146, 154, 161, 179, 182, 184, 190,

191, 197, 205, 209, 210, 232

U.S. Environmental Protection Agency....... 3, 4, 6, 13, 14,

22, 23, 24, 42, 43, 49, 55, 59, 63, 64, 65, 66, 67,

68, 70, 74, 75, 76, 77, 124, 139, 140, 141, 142,

143, 147, 164, 167, 184, 191, 192, 193, 197, 198,

202, 210, 213, 218, 234, 234

U.S. Food and Drug Administration ....................... 170, 234

U.S. Geological Survey ........... 55, 63, 78, 147, 184, 187,

202, 236

underground storage tanks ............ 62, 63, 67, 68, 76, 236

ungulates .......................................................................... 204

urban runoff ........................................... 65, 66, 67, 78, 79

urban-wildland intermix .................................................. 204

urbanization ......................................... 62, 63, 78, 79, 166

USGS ...................... 55, 63, 78, 147, 184, 187, 202, 236

UST ............................................... 62, 63, 67, 68, 76, 236

Vibrio ........................................................................ 23, 165

virus ..................................................... 22, 23, 64, 83, 140

VOC ................................................................. 70, 236, 241

volatile organic compound .............................. 70, 236, 241

wading birds ........................................................... 164, 241

wastewater .............. 20, 21, 32, 42, 43, 44, 49, 55, 58,

62, 63, 64, 65, 66, 67, 74, 76, 77, 79, 169, 170,

180, 182, 197, 198, 199

wastewater treatment ..... 10, 20, 21, 32, 42, 43, 44, 49,

62, 63, 65, 64, 66, 67, 77, 79, 237

water birds ................. 164, 166, 167, 183, 204, 205, 213

water chemistry ........................................ 31, 55, 113, 169

water diversions ................................................. 44, 59, 182

Water Erosion Prediction Project ............................. 90, 236

water recreation .................................................... 74, 75, 79

water treatment plants........... 14, 22, 32, 43, 46, 49, 50,

74, 124

water treatment technologies ............. 225, 226, 227, 228,

229, 233

watershed ...... 7, 17, 26, 30, 31, 32, 34, 35, 36, 37, 38,

39, 42, 43, 45, 47, 48, 49, 50, 55, 56, 62, 66, 67,

77, 79, 81, 82, 86, 87, 90, 92, 95, 96, 97, 103,

110, 111, 113, 120, 121, 122, 124, 129, 134, 146,

147, 153, 154, 155, 159, 160, 167, 184, 197, 198,

199, 200, 202, 209, 210, 211, 212, 213, 214, 218

Watershed and SedimentYield Model ...................... 90, 236

Watershed Erosion Prediction Model ....................... 90, 236

Watershed Protection Program ................................ 167, 236

WATSED .................................................................. 90, 236

weed control .................................................. 140, 141, 146

well ..... 24, 58, 60, 65, 69, 70, 78, 143, 144, 145, 146,

190, 191, 192, 193, 204, 237, 238, 240, 241

well head ................................................................... 60, 191

WEPP........................................................................ 90, 236

wilderness .............................. 33, 34, 39, 81, 82, 83, 155

wildfire ................. 36, 37, 124, 126, 128, 129, 132, 135

wildlife ...... 55, 56, 139, 154, 158, 159, 160, 161, 162,

183, 188, 192, 199, 204, 209, 213, 218

WPP ........................................................................ 167, 236

Yersinia ..................................................................... 23, 165

zebra mussel .................................................................... 172

Subject Index
Dissmeyer, George E., ed. 2000. Drinking water from forests and grasslands: a

synthesis of the scientific literature. Gen. Tech. Rep. SRS–39. Asheville, NC:

U.S. Department of Agriculture, Forest Service, Southern Research Station.

246 p.

This report reviews the scientific literature about the potential of common forest

and grassland management to introduce contaminants of concern to human health

into public drinking water sources. Effects of managing water, urbanization,

recreation, roads, timber, fire, pesticides, grazing, wildlife and fish habitat, and

mineral, oil, and gas resources on public drinking water source quality are reviewed. Gaps in knowledge and research needs are indicated. Managers of national

forests and grasslands and similar lands in other ownerships, environmental

regulators, and citizens interested in drinking water may use this report for assessing contamination risks associated with land uses.

Keywords: Economics, nutrients, pathogens, sediments, source water assessments,

toxic chemicals.
The Forest Service, U.S. Department of

Agriculture (USDA), is dedicated to the

principle of multiple use management of the

Nation’s forest resources for sustained yields of

wood, water, forage, wildlife, and recreation.

Through forestry research, cooperation with the States and

private forest owners, and management of the National Forests

and National Grasslands, it strives as directed by Congress to

provide increasingly greater service to a growing Nation.

The USDA prohibits discrimination in all its programs and

activities on the basis of race, color, national origin, sex, religion,

sex, political beliefs, sexual orientation, or marital or family

status. (Not all prohibited bases apply to all programs). Persons

with disabilities who require alternative means for communication of program information (Braille, large print, audiotape, etc.)

should contact the USDA’s TARGET Center at 202-720-2600

(voice and TDD).

To file a complaint of discrimination, write USDA, Director,

Office of Civil Rights, Room 326-W, Whitten Building, 1400

Independence Avenue, SW, Washington, DC 20250-9410 or call

202-720-5964 (voice and TDD). USDA is an equal opportunity

employer.