United States

Department of

Agriculture

Forest Service

Southern

Research Station

General Technical

Report SRS–50

Proceedings of a Conference on

Sustainability of Wetlands and

Water Resources:

How Well Can Riverine Wetlands Continue to

Support Society into the 21st Century?

Oxford, Mississippi

May 23–36, 2000
DISCLAIMER

The use of trade or firm names in this publication is for reader information and does not imply endorsement by the U.S.

Department of Agriculture of any product or service.

PESTICIDE PRECAUTIONARY STATEMENT

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.

Papers published in these proceedings were submitted by authors in electronic media. Editing was done to ensure a

consistent format. Authors are responsible for content and accuracy of their individual papers.

March 2002

Southern Research Station

P.O. Box 2680

Asheville, NC 28802

Cover: Aerial photograph of experimental ponds and mesocosms at The University

of Mississippi Field Station (western portion).

Photo by Kevin R. Pigott
i

Proceedings of a Conference on

Sustainability of Wetlands and Water Resources:

How Well Can Riverine Wetlands Continue to

Support Society into the 21st Century?

Edited by

Marjorie M. Holland, Melvin L. Warren, and John A. Stanturf

Oxford, Mississippi

May 23–25, 2000

Hosted by

The University of Mississippi

Sponsored by

The University of Mississippi

USDA Forest Service, Southern Research Station
ii
iii

Table of Contents

Page

Human Alterations of Earth’s Fresh Water:

Scale, Consequences, and a Call to Action ................ 1

Sandra Postel

Institutional Change and Campus Greening at

Tulane University .......................................................... 4

Aaron S. Allen

The Collaboration Between the University of

Mississippi and Belize Enables Opportunities for

Science Education ....................................................... 14

Richard A. Belisle

Feasibility of Using Ornamental Plants in Subsurface

Flow Wetlands for Domestic Wastewater Treatment .... 17

Marco A. Belmont and Chris D. Metcalfe

Rehabilitation of Coastal Wetland Forests Degraded

Through Their Conversion to Shrimp Farms ................ 20

Peter R. Burbridge and Daniel C. Hellin

Effects of Hydrologic Conditions on Biogeochemical

Processes and Organic Pollutant Degradation in

Salt Marsh Sediments .................................................. 30

W. James Catallo

Restoration Methods for Deepwater Swamps ............. 39

William H. Conner, Kenneth W. McLeod, and

Ellen Colodney

Wetlands Systems in Southern Thailand:

The Essential Resources for Sustainable

Regional Development ................................................ 43

Rotchanatch Darnsawasdi and Prassert Chitpong

The Development of a Decision Support System

for Prioritizing Forested Wetland Restoration Areas

in the Lower Yazoo River Basin, Mississippi ................ 47

Angela A. Davis, Barbara A. Kleiss,

Charles G. O’Hara, and Jennifer S. Derby

States Assuming Responsibility Over Wetlands:

State Assumption as a Regulatory Option for

Protection of Wetlands ................................................. 57

Kristen M. Fletcher

Little Cypress Creek Study:

A Watershed Restoration Case Study ......................... 65

Lisa Gandy, Randy Roberson, and Tom Foti

Bottomland Hardwood Afforestation:

State of the Art .............................................................. 75

Emile S. Gardiner, D. Ramsey Russell,

Mark Oliver, and Lamar C. Dorris, Jr.

Page

Wetland Education through Cooperative Programs

Between Coastal Carolina University and

Horry County Public Schools ....................................... 87

Sharon L. Gilman

Wetlands and Agriculture: Are We Heading for

Confrontation or Conservation ..................................... 88

Brij Gopal

Three Mississippi Ecotypes of Wetland Plants ........... 94

Janet M. Grabowski

Winter Bird Communities in Afforestation:

Should We Speed Up or Slow Down

Ecological Succession? ............................................... 98

Paul B. Hamel, Daniel J. Twedt, Timothy J. Nuttle,

Christopher A. Woodson, Fred Broerman, and

Joseph M. Wahome

Globe, Student Inquiry, and Learning Communities .... 109

C.L. Henzel

Using Kepone to Exemplify the Importance of

Natural Variability in Estimating Exposure to

Toxic Chemicals from Aquatic Environments ............... 112

Robert J. Huggett

Potential Effects of Restoration on Biogeochemical

Functions of Bottomland Hardwood Ecosystems ........ 116

Graeme Lockaby and John Stanturf

Forest Restoration in the Nordic Countries .................. 120

Palle Madsen, Ása L. Arad•ttir, Emile Gardiner,

Pelle Gemmel, Kåre Lund Høie, Magnus Löf,

John Stanturf, Peter Tigerstedt, Hardi Tullus,

Sauli Valkonen, and Veiko Uri

Application of the Soil Perturbation Index to

Evaluate Created and Restored Wetlands .................. 126

Rebecca Smith Maul and Marjorie M. Holland

Pesticide Mitigation Capacities of

Constructed Wetlands .................................................. 133

Matthew T. Moore, Charles M. Cooper,

Sammie Smith, Jr., and John H. Rodgers, Jr.

Nutrient Storage Rates in a Natural Marsh

Receiving Waste Water ................................................ 135

J. A. Nyman

Effects of Soil Oxidation-Reduction Conditons on

Internal Oxygen Transport, Root Aeration, and

Growth of Wetland Plants ............................................. 139

S.R. Pezeshki and R.D. DeLaune
iv

Page

Riverbank Restoration in the Southern United States:

The Effects of Soil Texture and Moisture Regime on

Survival and Growth of Willow Posts ............................ 146

S. Reza Pezeshki, Steven D. Schaff, and

F. Douglas Shields, Jr.

Differences in Net Primary Production and

Biogeochemistry Between Contrasting Floodplain

Forests .......................................................................... 153

Erik B. Schilling and B. Graeme Lockaby

Use of a Constructed Wetland to Reduce

Nonpoint-source Pesticide Contamination of the

Lourens River, South America ..................................... 154

Ralf Schulz

Forest Restoration in a Global Context ........................ 160

John A. Stanturf

Page

Forest Linkages to Diversity and Abundance

in Lowland Stream Fish Communities ......................... 168

Melvin L. Warren, Jr., Wendell R. Haag,

and Susan B. Adams

Origins, Fates, and Ramifications of

Natural Organic Compounds of Wetlands ................... 183

Robert G. Wetzel

Development of a Short-term (<12 days),

Plant-based Screening Method to Assess the

Bioavailability, Bioconcentration, and Phytotoxicity

of Hexahydro-1,3,5- Trinitro-1,3,5-Triazine (RDZ) to

Terrestrial Plants ........................................................... 190

Linda Winfield, Steven D’Surney, and

John Rodgers
v

ACKNOWLEDGMENTS

We would like to thank all of the organizations that were cosponsors of the conference: The Croft Institute for International

Studies at The University of Mississippi; The Fertilizer Institute; The Henry L. & Grace Doherty Charitable Foundation, USDAAgriculture Research Service, National Sedimentation Laboratory; USDA Forest Service, Department of Bottomland Hardwood Research; The Phil Hardin Foundation; The National Center for Natural Products Research; and The National Science

Foundation.

We would like to thank all of the organizations that made contributions to the conference: Department of Biology, The University of Mississippi; Fisher Scientific; Sigma Xi, The University of Mississippi Chapter; Society of Wetland Scientists (SWS),

National Board; Society of Wetland Scientists, South Central Chapter; USDA Natural Resources Conservation Service, Wildlife

Habitat Management Institute; USDA Natural Resources Conservation Service, Wetland Science Institute.

We would like to thank all of the organizations that endorsed the conference: International Association of Theoretical and

Applied Limnology; The Ecological Society of America; The Lower Mississippi River Conservation Committee; The Mississippi

River Foundation; American Fisheries Society; Society of Environmental Toxicology and Chemistry (SETAC).

We would like to thank the Conference Organizing Committee: Dr. David Arnold, Mississippi Chemical Corporation; Dr.

Elizabeth Blood, Jones Ecological Research Center; Dr. W. Franklin Harris, Oak Ridge National Laboratory; Dr. Marjorie M.

Holland, The University of Mississippi; Dr. Michael Metcalf, Croft Institute for International Studies; Dr. Gary Miller, The University of Mississippi; Dr. Andy Mullins, The University of Mississippi; Ms. Kathy Sukanek, The University of Mississippi.

We would like to thank the members of the Conference Planning Committee: Dr. Nolan Aughenbaugh, The University of

Mississippi; Dr. William Benson, U.S. Environmental Protection Agency; Dr. Alice Clark, NCNPR, The University of Mississippi;

Dr. Charles Cooper, USDA-ARS National Sedimentation Laboratory; Dr. Gary Gaston, The University of Mississippi; Dr.

Marjorie M. Holland, Chair, The University of Mississippi; Mr. Gerald Inmon, USDA Forest Service (retired); Ms. Janet Jenkins,

Bruce High School; Dr. Sammy King, U.S. Geological Survey; Dr. Barbara Kleiss, U.S. Geological Survey; Dr. Al Mikell, The

University of Mississippi; Dr. Reza Pezeshki, The University of Memphis; Dr. Catherine Pringle, The University of Georgia; Mr.

Paul Rodrigue, USDA-NRCS; Dr. Matt Romkens, USDA-ARS, National Sedimentation Laboratory; Dr. Larry Shaffer, The

University of Mississippi; Mr. Duane Shaw, MDWFP; Dr. Jean Shaw, The University of Mississippi; Dr. Stephen Threlkeld, The

University of Mississippi; Dr. Melvin Warren, USDA Forest Service; Dr. Robert Wetzel, The University of Alabama; Mr. Darryail

Whittington, MS DEQ; Dr. Robert Woolsey, The University of Mississippi; Dr. Joy Zedler, The University of Wisconsin.

Last, but certainly not least, special thanks go to Keith Wright, a student assistant with the UM Field Station, who devoted

considerable talent and skill to formatting and reviewing the proceedings manuscripts.
vi
1

HUMAN ALTERATIONS OF EARTH’S FRESH WATER: SCALE,

CONSEQUENCES, AND A CALL TO ACTION

Sandra Postel1

Abstract—One of the biggest challenges society faces in this new century is figuring out how to satisfy the water demands

of 8 billion people while at the same time protecting the aquatic ecosystems and ecological services that humans and all

species depend upon. Since 1950, water demands worldwide have more than tripled, while the scale of our dams and

reservoirs, river diversions, and groundwater exploitation have fundamentally altered hydrological systems and the

ecological services they perform. On top of all these stresses, we face the added complexity of global climate change. We

are going to need a fundamentally different approach if we are going to have any hope of maintaining some degree of

ecological health in the face of these rising demographic and economic pressures. This paper addresses some of the goals

and key policy levers necessary for wetland ecosystem restoration and protection. This paper is a transcription of Sandra

Postel’s plenary talk at the May 2000 Conference.

1

 Director, Global Water Policy Project, 107 Larkspur Dr., Amherst, MA 01002-3440

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability

of wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.

INTRODUCTION

One of the biggest challenges society faces in this new

century is figuring out how to satisfy the water demands of

eight billion people while at the same time protecting the

aquatic ecosystems and ecological services that humans

and all species depend upon. This paper explains why I think

this is the case, and explores the critical role of science and

scientists in helping meet this challenge.

We have an unsettling degree of uncertainty when it comes

to just about all aspects of the world’s water—from the

natural hydrological cycle itself to the amount of water we

have stored in reservoirs, to how much groundwater is stored

in aquifers, to how much water humanity currently uses. Most

of the numbers discussed below are best estimates rather

than precise facts. There’s no doubt that we need a much

better scientific understanding of many aspects of

freshwater. But we know enough to say unequivocally that

over the last century, human activities have altered the

hydrologic cycle and hydrologic regimes in unprecedented

ways and on a global scale—and the costs and

consequences of these impacts have just begun to come to

light.

Since 1950, water demands worldwide have more than

tripled. We now remove from the earth’s rivers, streams,

lakes, and aquifers about 4,000 cubic kilometers of water per

year. Irrigated agriculture has been by far and away the

biggest driver behind this rise in water use. Irrigated land

worldwide has nearly tripled over the last half century,

climbing from 100 million hectares to more than 270 million

hectares today. Agriculture now accounts for about 70

percent of world water use, industries for about 20 percent,

and cities and towns for about 10 percent.

At first glance, it might seem that at current levels of water

use we’re still in reasonably good shape. Our current use of

4,000 cubic kilometers represents only about 10 percent of

the world’s total renewable runoff—the water that annually

flows back toward the sea via rivers and aquifers. But if we

look more carefully at how much of that river and

groundwater flow is actually accessible to us—and therefore

can serve as a supply for agriculture, industry, and cities—

we find that the situation isn’t at all comfortable. Only about

31 percent of global runoff can be tapped where and when

we can use it. And of this, we already appropriate about half

to meet current human needs—which suggests that global

limits are closer than we’d previously realized. The only way

to increase the accessible runoff is to capture and store

more flood water, which typically means constructing more

dams. So we face a kind of Faustian bargain with Nature: In

order to meet future human needs, it seems, we need to

shift an even larger portion of the world’s water from serving

Nature’s purposes to serving humanity’s purposes. And I will

come back to this dilemma.

But first, a few reality checks. If we are anywhere close to

hitting global limits, we should be seeing physical signs of

water stress and unsustainable uses in many freshwater

systems and across fairly wide geographic regions. So let’s

look at what is happening.

Rivers

We have seen the number of large dams (those at least 15

meters high) climb from 5,000 in 1950 to more than 40,000

today. We have a less good count of small dams, but they

number somewhere in the 800,000 range. The reservoirs

behind these dams are capable of storing nearly 20 percent

of total annual global runoff, and as is often the case, this

global figure masks great variation and extremes among

individual river systems. Lake Nassar, behind Egypt’s Aswan

Dam, is able to fully store two years worth of the Nile’s flow,

for example.

Recent surveys suggest that nearly 60 percent of the world’s

largest 237 rivers are moderately to strongly fragmented by
2

dams and diversions. Perhaps more dramatically, we’re now

seeing that many major rivers are tapped out during the dry

season, with virtually no freshwater reaching the sea for

months at a time. These include five of Asia’s great rivers—

the Ganges and Indus Rivers in South Asia, the Amu Darya

and Syr Darya Rivers in Central Asia, and the Yellow River in

China. The Yellow River first ran dry in 1972. But the

situation has worsened greatly over the last 15 years. I

remember traveling by train in China in 1988, and awakening

early so as not to miss crossing the Yellow River, and being

stunned by how little water this great river—the cradle of

Chinese civilization-was actually carrying. In 1997 the Yellow

ran dry for a record 226 days. Closer to home, the Colorado

River doesn’t reach the sea at all in an average year, and

that’s been true more or less since completion of Glen

Canyon Dam in the early 1960s.

Groundwater

Groundwater, which forms the base flow of rivers, has come

under heavy pressure in the last 50 years as well. Prior to

WWII, we just didn’t have the technical capabilities of

pumping groundwater on a vast scale. But with the advent of

powerful pumps and deep drilling technologies, farmers and

other water users began to tap groundwater on a historically

unprecedented scale. In India, the number of groundwater

wells climbed from 4 million in 1951 to 17 million in 1997. In

China, the number of irrigation wells has climbed 20-fold

over the last three decades. Here in the US, farmers drilled

millions of wells into the Ogallala Aquifer—one of the

planet’s greatest underground water reserves, and which

now waters 20 percent of U.S. irrigated land.

The problem is that much of this groundwater use is not

sustainable. In just about every area of intensive

groundwater-based agriculture, water tables are dropping

steadily because pumping exceeds recharge. This is the

case in the Punjab of India, the north plain of China, the U.S.

Great Plains, California’s Central Valley, and much of North

Africa and the Middle East. This is a big red flag for future

food security, since as much as 5-10 percent of the world’s

food production may now depend on the overpumping of

groundwater. And it’s also a great ecological threat as the

lowering of water tables dries up springs and wetlands—

and, in some cases, turns perennial rivers into seasonal

ones by eliminating their base flow.

Together, the scale of our dams and reservoirs, river

diversions, and groundwater exploitation have fundamentally

altered hydrological systems and the ecological services

they perform. It’s estimated that half of the world’s wetlands

have been lost during the 20th century—in part due to direct

conversion to agricultural and urban land uses, but also due

to the cutting off of rivers from their flood plains and their

deltas, and to the overpumping of groundwater.

In many countries, 10-30 percent of freshwater fish are now

threatened with extinction. Here in the United States, The

Nature Conservancy reports that 38 percent of freshwater

fish are at risk, along with 69 percent of freshwater mussels,

and 51 percent of crayfish. The U.S. ranks first in the world in

the number of known species of freshwater mussels and

crayfish and 7th in the number of known freshwater fish

species. Freshwater species are proportionately at greater

risk than others, and the principal reason for their

imperilment is the destruction and degradation of aquatic

habitats.

We could, if we had time, run down a long list of aquatic

ecosystems on every continent that are in states of rapid

decline—from the most dramatic case of the Aral Sea in

central Asia, which used to be the world’s fourth largest lake,

to the Everglades of south Florida, the Danube Delta of

Europe, the Ganges and Indus deltas of south Asia, to entire

river systems like the Colorado, the Missouri, the Nile, and

the Rhine.

On top of all these stresses, we face the added complexity

of global climate change. The patterns of river runoff we

have measured during the 20th century will almost certainly

not be good guides for our planning in the 21st century.

Rainfall patterns are expected to shift and droughts and

floods to intensify. Many of the world’s major rivers are fed by

mountain snowpacks, which are fantastic natural reservoirs.

They store water through the winter and release it during the

spring and summer. As temperatures warm, more of that

winter precipitation will fall as rain rather than snow, and the

snowpack will melt earlier and faster. In many places, this

will mean more intense flooding in the early spring and lower

flows during the summer, when water demands are highest.

So we have a large backlog of water problems to confront,

even as pressures on water systems continue to increase.

Just to supply the food demands of 2025, we could need to

find an additional 800 cubic kilometers of irrigation water - a

volume nearly equal to the annual flow of ten Nile Rivers.

So I return to our Faustian bargain: We can continue to

extract more water from Nature in order to meet rising water

demands, but in doing so we place in jeopardy the survival

of major pieces of the aquatic world, pieces we depend on

for a host of ecological services worth—no one knows how

much—but almost certainly in the hundreds of billions if not

trillions of dollars annually.

I would submit that there is no winning scenario we can

create out of these current trends in water use and

management. We are going to need a fundamentally

different approach if we are going to have any hope of

maintaining some degree of ecological health in the face of

these rising demographic and economic pressures.

What are some of the big priorities?

First, I believe we need a multi-disciplinary and crossprofessional effort to systematically determine the quantity,

quality, and timing of flows needed for freshwater

ecosystems to sustain their critical functions. Without an

effort such as this, no country or region can answer such

basic questions as: how much water is available to meet

human needs in a sustainable fashion? Or, how must dams

be operated in order to sustain critical ecosystem functions

and to protect biodiversity? This initiative would have to be

scientifically credible but policy-focussed-that is, produce a

strategy that can be implemented. South Africa is the only

country I know of that has adopted this kind of goal as a

matter of national policy. South Africa’s new water policy

states that, “The quantity, quality, and reliability of water
3

required to maintain the ecological functions on which

humans depend should be reserved so that the human use

of water does not individually or cumulatively compromise

the long term sustainability of aquatic and associated

ecosystems.” Many of the aquatic scientists in South Africa

are now involved in the effort to determine these

“environmental water reserves” in some way. Of course it

remains to be seen how effectively this policy will be carried

out.

Another bold step has been taken in Australia’s MurrayDarling river basin, where the Commission has placed a Cap

on total water extractions from the basin. The rationale here

is simply that the river system has already sustained too

much ecological harm, and so no additional water diversions

should take place. New demands for water in the region are

to be met through conservation and water trading rather

than by increasing the supply.

Here in the United States, we have had several calls for

major initiatives to generate the knowledge-base needed for

sound freshwater management and protection of

ecosystems. Bob Naiman, Diane McKnight, Jim Karr and

others wrote eloquently of the need for a national initiative of

this sort in an article in Science in 1995. These and other

researchers have written extensively about the geomorphic,

hydrologic, and biological parameters that would form the

core of this knowledge base for freshwater ecosystem

protection and restoration.

I think the time is right to push harder for an initiative such

as this, because much has changed in the last five years.

When the Edwards Dam on the Kennebec River in Maine

was demolished in the summer of 1999, it was a very public

symbol and signal—and Secretary Babbitt said this—that

dams are not forever. The idea that we’d be having a serious

discussion about breaching major dams in the Columbia

River basin—four on the lower Snake—would probably have

seemed outlandish even 5-10 years ago.

These debates signal a major shift in public attitudes that

provides new opportunities. Many more small dams will likely

come down over the next few decades, but probably more

important is the possibility of rethinking the operation of

dams that will remain standing. To what degree can we

mimic a river’s natural hydrograph and still meet a portion of

the economic uses for which a given dam was built? What

will be the ecological benefits of doing this? Marginal

tradeoffs like this will be made, and these collectively may

be more important than the big-ticket items like a few big

dams coming down.

What are the key policy levers for ecosystem restoration and

protection that we’re likely to see greater use of?

• FERC relicensing: of private hydro dams (e.g., Edwards

Dam)

• Endangered Species Act—which, if invoked in its full

force, would dramatically alter water use and river

management nationwide.

• Instream flow requirements—On a case by case basis,

these are beginning to get clarified. Judges are making

decisions about what “impairment” of a river system

means in order to determine what uses and extractions of

the river are reasonable and acceptable. This is not just

happening in the West, but in the East as well. Very

interesting case of the Shepaug River in western

Connecticut decided in February 2000.

• Public Trust Doctrine—applied in a historically new and

much broader way in the case involving Mono Lake in

California. It is not yet clear whether this will be a

precedent for broader applications.

Unless we have the knowledge base to assist in these legal,

regulatory, and management debates, and are able and

willing to translate that knowledge base to policymakers and

water managers, we will miss big opportunities for

conservation and restoration of aquatic ecosystems.

Second, we need an all-out effort in every sector of the

economy to raise water productivity. We simply cannot hope

to achieve our ecosystem protection and restoration goals if

we do not promote more efficient use of water. I believe that

with an all-out effort to promote more efficient and equitable

use of water worldwide, we could satisfy year-2025 water

demands without extracting much more water from the

natural environment than we do today. That’s an exciting

prospect-but we’re a long way from getting there.

I once heard it said that “Technology is Nature’s experiment

with Man.” We need to get smarter about the technologies

we employ and how we employ them—whether it’s how we

control floods, how we irrigate our crops, or how we operate

our dams.

To do this we need to build new research and

management partnerships, especially ones that bridge

ecology, engineering, and economics. Some of the best

attempts I’ve seen to evaluate ecosystem services have

been done through interdisciplinary and/or crossprofessional collaborations. We need more efforts not just

to think out of our box, but to come out of our box. Even

applied science is not getting used in making good policy

and management decisions because too few scientists

are bridging the divide to the policy world.

Finally, I also believe that, as scientists and citizens, we can

contribute more to the resolution of our water problems by

advocating more forcefully for the adoption of a guiding

water ethic in society. The fact that water is the basis of life

lends an ethical dimension to every decision that’s made

about how to use and manage it. Especially in the face of

scientific uncertainty and potentially irreversible change,

appealing to an ethic of protection and preservation of

freshwater ecosystems can be both justified and compelling.

Eleanor Roosevelt once said, “We should constantly be

reminded of what we owe in return for what we have.” I

would submit that what we have is more knowledge about

the health of our aquatic ecosystems and the threats to them

than the vast majority of people on the planet, and that what

we owe to society in return is the expression of our best

judgement—not perfect judgement, but best judgement—of

how to preserve and protect these critical ecological assets.
4

INTRODUCTION

“Greening the campus” means increasing environmental

awareness or action or both on campus—in the operational

facilities and processes of the campus as well as in the

human communities of the campus and surrounding areas.

Greening the campus involves working towards some or all

of the goals set forth in the Blueprint for a Green Campus.3

Although the fundamental theme of greening is education,

this study focuses on campus operations, the greening of

which is pedagogical process itself.

The economics of campus environmental initiatives in higher

education are well documented: greening the campus saves

money. Twenty-three conservation initiatives at fifteen U.S.

institutions of higher education each saved between $1,000

and $9 million, with total annual savings at $16.8 million

(Eagan and Keniry 1998).4 Investing in campus greening is

therefore an economic, educational, and environmental

investment with handsome returns—both financial and social.

In addition to saving money, campus greening allows

students to learn how to infuse environmental sustainability

into the larger society. Students must be able to practice

(and see the university practice) the lessons of

environmental sustainability, which they are taught in the

classroom. Tulane has committed to environmental studies

along with three other areas of interdisciplinary interest:

urban studies, international studies, and information

technology. Together, the four are conducive to

environmental responsibility and stewardship.

Tulane is located in New Orleans, LA, the southernmost port

on the 2,552-mile-long Mississippi River. The Mississippi

River Basin drains 41 percent of the landmass of the

continental United States. The river is the dominant feature

of New Orleans, and the university has designed research

agendas and teaching curricula around it. Tulane was

established in 1834 as the Medical University of Louisiana to

study and treat “the peculiar diseases which prevail in this

part of the Union” (Tulane University 1997). The university is

now comprised of 11 academic divisions with approximately

6,500 undergraduates, 4,800 graduate students, and 8,000

employees, of which approximately 1,750 are full- or parttime faculty. Tulane is responsible for approximately 24,000

jobs in Louisiana and an annual injection of nearly $1.5

billion into the local economy (Strecker 1998). With its

location on the Mississippi River, traditional focus on health,

and impact on the local economy, Tulane has a formidable

presence in the Southern United States.

INSTITUTIONAL CHANGE

A Model for Institutional Change

Figure 1 is the model of institutional change. It is derived

from a literature review of institutional change in higher

education.5 Additionally, case studies in nonenvironmental

INSTITUTIONAL CHANGE AND CAMPUS GREENING

AT TULANE UNIVERSITY

Aaron S. Allen1

1 Environmental Studies Program, 201 Alcee Fortier Hall, Tulane University, New Orleans, LA 70118.

2 This article is excerpted from a larger study by the author, Greening the Campus: Institutional Environmental Change at Tulane University

(1999). It is available on the Internet at www.tulane.edu/~env_stud/greening.htm .

3

 The Blueprint (1995) mentions ten items: integrating environmental knowledge into all relevant disciplines; improving environmental course

offerings; providing opportunities to study campus and local environmental issues; conducting a campus environmental audit; purchasing

environmentally responsible products; reducing campus waste; maximizing energy efficiency; making environmental sustainability a top priority

in campus land-use, transportation, and building planning; establishing a student environmental center; and supporting students who seek

environmentally responsible careers.

4

 The summary of financial data from Eagan and Keniry (1998) is available at www.nwf.org/nwf/campus/tools/publications/gigr/cost.html .

5

 The complete literature review is in the Greening the Campus study.

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability of

wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.

Abstract—A case study of Tulane University that examines the institutional change process is presented in this paper.

Agents of change can use the examples and conclusions as a basis for making changes at any institution. The inability for

Tulane to make the campus environmentally sustainable in terms of operations and education was due to the lack of an

institutionalized internal lobbyist and leader dedicated to environmental issues. That argument is supported with a model for

institutional change, a historical analysis of nonenvironmental and environmental change initiatives at Tulane, a review of

campus greening programs in institutions of higher education, and a series of interviews with Tulane students and employees. In the summer of 1999, as a result of an earlier version of this study,2 an Office of Environmental Affairs (OEA) was

created according to the “Blueprint for a Green Tulane,” which outlined the steps necessary for institutional environmental

change to occur. The central component of that change is leadership from the OEA’s Environmental Coordinator and from

students who will, in turn, carry their leadership beyond the campus to create a more sustainable world.
5

and environmental change at Tulane and in academia

support the model, which itself is not restricted to

environmental change in any way. The key element is a

leader who is an administrator or faculty member but not a

student because students lack power and connections and

are temporary. (Students, however, do play absolutely

integral roles in the change process, as discussed in a later

section.) In addition to the leader, leadership from the

administration is necessary to support the change agenda.

The model is a conceptual framework for understanding

and implementing change. It is dynamic: the dark arrows

represent normal “flow,” whereas open arrows represent

feedback. The model is dynamic not only in itself but also

between applications; different circumstances result in

different paths. For example, education (the “end”) may

result in further advocacy for new changes (thus the dotted

line, effectively making the model cyclical); also, procuring

policy may return the advocates to the advocacy stage

Figure 1—Schematic of the model for institutional change.
6

before getting resources. The model is not rigid; for

example, policy may be skipped entirely, but the results of

the change may not be permanent. Dividing the change

process into the segments of the model is artificial but

necessary; institutional change is not spontaneous, and

greater understanding of the process will increase the

likelihood of success for change movements.

Advocacy is the impetus to begin change. It is the product of

diffuse, irregular efforts of (primarily) students and faculty

found in the “shadow” of the university—the area outside of

the “mainstream” of campus life and separate from the

traditional governing structures of the institution (Bowers

1997, Mansfield 1998). Advocacy is usually a grassroots or

bottom-up effort, but top-down advocacy is just as important:

the two converge in the middle to create the integrated

advocacy required for institutional change.

Advocacy results in policy. Development of specific and

general policies should be consensual with the input of all

appropriate parties. Policies should be applicable,

enforceable, and nonrhetorical in order to support, justify, and

communicate the change goals. Additionally, policy

development and having policies in place are forms of

education (a means and end) about the change agenda

(Altbach 1974, Cerych and Sabatier 1986, Creighton 1998,

Fantini 1981, Gitell 1981, Hamburg and Ask 1992, Keniry

1995, Lane 1990, MacTaggart 1996, Smith 1993, Strauss

1996).

Advocacy and policy procure resources. Roughly prioritized,

the primary resources are personnel (a leader, support staff,

an office), financial resources, information and data, power

(or direct access to power), and the ability to offer

opportunities and incentives for improvement and positive

change. Resource allocations should be in line with the

missions of the institution, and a continual supply of

necessary resources will maintain the desired changes

(Altbach 1974, Cerych and Sabatier 1986, Creighton 1998,

Dominick 1990, Fantini 1981, Gitell 1981, Hamburg and Ask

1992, Keniry 1995, Lane 1990, MacTaggart 1996, Smith

1993, Strauss 1996).

Leadership is the key and defining element of the model for

institutional change. Advocacy procures the leader, who is

supported with policy and resources. The leader is in an

institutionalized position dedicated to the change agenda.

He or she is the change agent: the communicator and

facilitator of the change process, the advocate and lobbyist

for the change agenda. The leader needs power or direct

access to power. The institution—especially the

administration, which should also act as leaders for

change—must support the leader. Finally, the leader should

be charismatic. Important character traits include

communication, interpersonal and listening skills, visionary

planning, and the capability to accomplish meaningful

projects (Berry and Gordon 1993, Creighton 1998) (see also

Altbach 1974; Cerych and Sabatier 1986; Creighton 1998;

Dolence and Norris 1995; Dominick 1990; Fantini 1981;

Farmer 1990; Gitell 1981; Hamburg and Ask 1992; Keniry

1995; Lane 1990; MacTaggart 1996; Orr 1990, 1992, 1994,

1995, 1996; Rainsford 1990; Smith 1993; Strauss 1996;

Wood 1990).

While the leader is the key element to the model, it is also

the place for the tragic flaw: how one person can do

everything. Some solutions include having other leaders and/

or support staff, or, as discussed below, having a guiding

committee and involving students in the change process.

The leader develops well-defined means to achieve agreedupon ends. Neither the means nor the ends can be rigid.

Means are the implementation plans; they are many and

specific, and they address education and process reengineering (physical and administrative). Ends are goals;

they are few and broad in scope. Examples of ends might be

ecological literacy of graduates and an environmentally

sustainable campus (Alinsky 1971) (see also Altbach 1974,

Dolence and Norris 1995, Eagan and Orr 1992, Farmer

1990, Keniry 1995, Lane 1990, Smith 1993, Wood 1990).

Education is the primary means and end. Campus decisionmakers must be educated on the mechanics of the means

and ends of the change agenda. The same issues should be

communicated to the entire campus since education about

the change agenda is not spontaneous, e.g., the campus

community must be educated on the mechanics of a

recycling program or the larger goals of environmental

sustainability. Eventually, the education reaches society, and

such is the ultimate goal for environmental change in higher

education (Ackerman 1997; Altbach 1974; Brown and Duguid

1996; De Young 1986; Dolence and Norris 1995; Gitell 1981;

Keniry 1995; MacTaggart 1996; Orr 1992, 1994; Smith 1993).

Some theory ties together the model for institutional change

in higher education. Change does not happen spontaneously

(Ackerman 1997, Bowers 1997, De Young 1986, Williams

1991). The changes pursued must be realistic. They will take

time to achieve and will never be 100 percent complete

(Cerych and Sabatier 1986, MacTaggart 1996, Steeples

1990). Operational changes affect some people significantly,

whereas most are affected only minimally; transformation,

not revolution, is needed. A two-dimensional framework of

change is appropriate for Tulane: depth is the degree to

which a change requires a departure from existing values

and practices, and breadth is the number of areas within the

institution that a change is expected to introduce

modifications. Wide/deep changes result in opposition, and

narrow/shallow changes do not take hold. Changes are most

likely to succeed when they are moderate in depth and

breadth of change (Cerych and Sabatier 1986). Institutional

environmental change with regard to campus operations is

moderate change (Hamburg and Ask 1992).

Nonenvironmental and Environmental

Institutional Change at Tulane

Six case studies of change initiatives at Tulane show that

moderate and profound changes are possible—given an

empowered leader (or leaders) with resources and policy

who introduces means and ends to implement change.

Multicultural Affairs, Bisexual, Gay, and Lesbian Affairs, and

Tulane College Programming show that it is necessary to

establish offices responsible for oversight and

implementation of changes. Advocacy began the

establishment of all three, and all established policies and

procured resources; then institutionalized leaders
7

implemented educational programs (means) to achieve

broad goals (ends). Two other reforms were more ambitious

in their scope: Tulane 2000 sought to stabilize the

university’s budget (and subsequently focus the institution’s

academic priorities) with cutbacks and resource

reallocations; the University Transformation Program sought

to improve the quality of staff services and classrooms,

create an extracurricular program for first-year students,

institute an information technology helpdesk, and establish

an international studies office. Both initiatives had a leader

(the president and the provost, respectively) and resources

to develop and implement policy to affect change. People did

not immediately embrace these issues (they were not

spontaneous); advocates and leaders convinced the campus

that they were meaningful changes. For example, Tulane’s

management takeover of the Housing Authority of New

Orleans was not spontaneous—the leader who initiated the

project believed that Tulane’s involvement was appropriate

and in the best interests of HANO, Tulane, and the citizens

of New Orleans.

The necessary elements of achieving change characterize

these preceding examples, and most fit into the strategic

goals of the university (urban studies, international studies,

information technology, and environmental studies). Missing

from these six initiatives, however, is a concerted effort to

make Tulane more environmentally responsible. Whereas

environmental research and education have improved

(largely due to grant funding), the third and critical element

of a green institution of higher education—operations— has

not been greened. The three divisions of the university are

research, education, and operations, and, at Tulane, each

has been greened to some extent.

Environmental research has been the most successful

division. It is a popular area because of the income

associated with research grants and the opportunities for

publishing. Also, quasi-policy (the environmental studies

focus) and resources (multimillion dollar grants) led the

development of extensive environmental research programs.

The leadership of Dr. John McLachlan of the Center for

Bioenvironmental Research at Tulane and Xavier

Universities (CBR) has developed, coordinated, and

maintained environmental research program opportunities.

The research division received a subjective grade6 of “A-” in

the spirit of the Green Gradecard for the Green Wave

environmental audit (discussed in a later section).

Tulane’s Environmental Studies Program (ENST) has a

history that epitomizes how institutional change occurs. In

the early 1970s, students lobbied for the creation of the

ENST, but the resulting coordinate major program (in which

students major in another field in addition to environmental

studies) stagnated until the early 1990s because the

program was not allocated a budget and had only the

devotion of one professor, who was not compensated for his

involvement. As a result of the then new environmental

studies focus of the university, the program progressed: new

faculty became involved and established an environmental

education committee, and grant monies provided the

resources to offer course development grants, purchase

equipment, hold training seminars,7 and hire a part-time

program coordinator. As a result, the program prospered,

and enrollment increased dramatically. But the faculty

leading the program could not dedicate enough of their

professional time to the program; they treated it as if it were

a University Senate committee. The student environmental

organization, the Green Club, worked cooperatively with the

ENST on numerous projects, including the Environmental

Forum Newsletter, campus environmental email lists, and

the design and publication of the Enviro Counter Culture

Catalog, a guide to environmental classes at Tulane, which

has received wide acclaim from within and outside of the

university.8 In 1998, the grants ended and the university did

not provide a budget for the ENST and its more than 50

students. The CBR stepped in to fund the program, but that

funding is also from grants. The future of the environmental

education program at Tulane is in question because of the

lack of institutional support (a budget). The program is still

directed by faculty members who receive no compensation

or official credit for their time. Whereas the ENST has

potential to be a top program at Tulane and in the Southern

United States, the lack of support and the absence of a fulltime dedicated leader are hindering such success. The

education division received the subjective grade of “B-”.

The Green Club and the Tulane Environmental Project

(TEP) have been significantly involved in the greening of

one operational aspect of Tulane: recycling. Recycling at

Tulane began in the 1970s as a volunteer effort. In the late

1980s, the Green Club formed to address more

institutionalized recycling. In the early 1990s, Green Club

leadership petitioned the university to establish a committee

to green the campus. Tulane’s president at the time, Dr.

Eamon Kelly, established the TEP and appointed the chair,

Oliver Houck, Professor of Environmental Law. The TEP was

active for 2 years. In the first year, the members of the TEP

researched and implemented a recycling program, hiring a

full-time coordinator and receiving a minimal university

budget. (In their second year, they began a recycled

procurement program to “close the loop,” but that initiative

was limited to a few paper products.) Peaks and troughs in

student leadership and activism (advocacy), the coming and

going of numerous recycling coordinators over the years

(leadership), and variable administrative support

(resources) have led to peaks and troughs in the success of

recycling operations. The Green Club has attempted other

operational greening programs, e.g., a “Green Dining”

initiative in Tulane dining areas, with minimal success. The

administration took on an economics-based lighting retrofit,

which did not include any education initiatives for saving

energy and had no explicit environmental motives, but no

other significant environmental initiatives have been

6

 The Green Gradecard did not use any standardized grading

procedure; the approximately 45 students who conducted the audit

relied on collective, subjective judgement to grade each area. The

same subjectivity was used in this study, although the research

behind the decisions was more extensive.

7

 Information on the faculty enrichment seminars is at

www.tulane.edu/~efes/ .

8

 The Catalog is available at www.tulane.edu/~greenclb/catalog/ .
8

institutionalized. The operations division received a

subjective grade of “D-”/“D”.

The history of greening at Tulane supports the model and

reaffirms the need for a leader. Research has had a

supported leader, and that division has been successful. As

for education, the Environmental Studies Program should be

a university-supported program with a leader. Recycling and

procurement programs are in need of improvement; each

needs policy, resources, leaders, and comprehensive means

to achieve those ends. Additionally, other campus greening

programs for operations need to be established for Tulane to

live up to its reputation as an environmental (research and

education) university.

THE GREENING PHENOMENON IN HIGHER

EDUCATION

The Green Gradecard for the Green Wave environmental

audit highlights many areas that are in need of improvement

at Tulane, especially when compared with other institutions

of higher education.9 Experiences in academia offer

caveats, lessons learned, and examples on which Tulane

can build and even exceed. The greening initiatives in

academia support the model for change, and they show the

sound economic, social, and environmental implications of

such programs (Blueprint 1995, Creighton 1998, Eagan and

Keniry 1998, Keniry 1995, Smith 1993).

Environmental audits are powerful tools for gathering

information about the environmental quality of the campus.

They are the starting point for environmental change, and

they provide information to educate the campus, the

community, and especially those involved in the audit.

Tulane’s audit, the Green Gradecard for the Green Wave,

which an Environmental Sociology class conducted in the

spring of 1997, issued letter grades with respect to various

areas of environmental performance. Environmental Studies,

an energy-saving lighting program, and hazardous waste

policies received “A-” grades, whereas recycling, investment

practices, and procurement of chemicals and pesticides

received failing grades. Overall, the audit graded 22 areas,

and Tulane’s “Green GPA” came out to a 1.9/4.0, or a “C”

average. The audit concluded that the university should

make an “institutional commitment to incorporate

environmental decision making into all facets of [campus]

operation . . . [and] establish a standing University Committee

for Environmental Affairs.” The Gradecard supports the

model for change in that it advocates institutional policy and

resources that would allow for administrative (leadership)

efforts to implement environmental change.

Programs at other institutions concerned with environmental

curricula and campus environmental consciousness illustrate

the essential role of leadership to provide education. Their

success is reflected in campus environmental cognizance.

Progressive environmental building, land use, and

transportation (parking) policies have social, administrative,

and economic benefits. Energy and water conservation

programs are financially sound and serve as education

about the importance of conserving natural resources. The

greening of food service operations has health,

environmental, and economic benefits for the campus and

local community. Waste issues (recycling, hazardous waste,

and medical waste) are visible to many in and out of the

campus community; greening them is fiscally responsible, is

educational, has positive impacts for the environment, and

improves the image of the institution. Green procurement

provides market stimulation to keep recycling and waste

reduction initiatives available and economical. Finally,

environmental research and socially responsible business

and investment practices have impacts that can be felt

around the world. Case studies from progressive and

innovative institutions in the above areas provide examples

of what and how Tulane can green (Creighton 1998, Eagan

and Orr 1992, Keniry 1995, Smith 1993). Many of the case

studies support the model for change.

HEARING FROM THE TULANE COMMUNITY

Interviews with Tulane students, staff, faculty, and

administrators further support the model. Five of the six

questions support the thesis of the study that a leader is

needed to institutionalize and carry out greening efforts.

The four main institutional change barriers, as determined

from the interviews, are institutional/organizational (lack of

communication, lack of advocacy, and the lack of a leader),

financial (lack of allocation of resources), cultural (lack of

education),10 and educational (lack of a modus for

education). Interviewees thought that greening programs

should relate to operations (administrative and physical) and

education (individual and community learning, both in and

out of the classroom). The results of the interviews clarify

roles of each tier of the university community: students as

learners, educators, and advocates; staff as learners and

empowered “doers”; faculty as advocates and educators

(who should practice environmental sustainability, especially

if they teach it); and administrators as leaders in all aspects

of the greening process. The responses for the roles of

administrators reiterated every element of the model and

focused on the need for an environmental coordinator to

lobby the administration on environmental issues. Finally,

interviewees affirmed that it is possible and appropriate to

green Tulane. It is now possible to formulate a proposal for

greening Tulane.

THE “BLUEPRINT FOR A GREEN TULANE”

The “Blueprint,” which is based on the model for change, is

the plan for implementing institutional environmental change

at Tulane. Included in it is the proposal to establish an Office

of Environmental Affairs (OEA) and create an Environmental

Coordinator position, both of which were accomplished in

the summer of 1999. The “Blueprint” is presented here as

outlined in the spring of 1999; changes in actual

implementation are presented in the Conclusion.

9

 Tulane’s mascot is the “Green Wave.” The Green Gradecard for the

Green Wave is available at www.tulane.edu/~greenclb/audit/

audit.html .

10 The “cultural barrier” is complex, and more research is necessary

to determine specific aspects that could be the target of educational

programs.
9

Advocacy

Re-establish/reinvigorate the Tulane Environmental Project

(TEP) as the Tulane Environmental Committee (TEC).

Ideally, President Cowen should initiate the new TEC; he

should also confirm all appointments to the TEC and

appoint the chair who would act as the presidential liaison.

The TEC would be charged with approving an annual

agenda for campus greening and reviewing OEA projects. A

working group from the TEC could develop such an agenda

and then continually work with the OEA. The Environmental

Coordinator of the OEA would report to the TEC, (which

would, in turn, answer to the president). The TEC would

meet once or twice each academic year with

representatives from the students (Associated Student Body

and Green Club), the staff (Staff Advisory Council), the

faculty (University Senate, CBR, and ENST), and the

administration (President’s Executive Working Group). Such

representation involves the research, education, and

operations divisions. The members of the TEC should be the

key players on campus with regards to environmental

change. The TEC would be the convergence of grassroots

advocacy, which has been displayed for years, and topdown advocacy, which has yet to be shown, while

simultaneously holding the power to make environmental

change. The TEC, the OEA, and the Environmental

Coordinator are interdisciplinary, interdivisional entities

pivotal for coordinating comprehensive institutional

greening.

Policy

Publish a statement that Tulane will be a leader in

environmental research, environmental education, and

environmental stewardship. The statement should outline the

core values of environmental responsibility that Tulane will

espouse. With such a proclamation, the TEC could gather

input from the university community via “town meetings” and

could draft a university environmental policy statement. The

president and the various legislative bodies of the university

could then ratify the policy. Additionally, it would be

necessary for the university to sign on to national and

international environmental platforms, e.g., the Talloires

Declaration and the Valdez Principles; such involvement

brings national and international attention as well as

assistance in implementing sustainability on campus. Finally,

project-specific policies, such as for recycling and

procurement, should be developed.

Resources

Seek funding for institutionalizing the OEA from internal and

external sources. Internal funds could first come from a

cooperative funding procedure, whereby each of the

academic deans, along with the vice presidents who would

be primary representatives on the TEC, would contribute

$3,000 to $5,000 for the job search and first year’s salary of

the Environmental Coordinator. With a job search estimated

at $3,500 to $4,000 and with salary and benefits estimated

at $36,000 to $36,500 (for a senior program coordinator

position), a total of approximately $40,000 is needed; with

eight academic deans and three vice presidents, the

cooperative funding program could work. (In the spring of

1999, no one was reluctant to contribute to such a

cooperative funding measure; they did mention that they

would be more willing to participate once they knew that the

president supported the OEA proposal.) This literal buy-in

into the OEA is important for developing cooperation.

External funds could come from alumni gifts and

endowments for programs, such as scholarships and

speaker series, and grants for projects and operating

expenses. An endowment of $1 million would secure the

OEA in perpetuity; the Office of Development could assist in

such fundraising. Some grants pending in the ENST include

such monies in anticipation of the OEA; the ENST has

found, however, that granting agencies will not pay for

employee salaries but are more likely to provide funds for

students, programs, and operating expenses. A study

sponsored by the Nathan Cummings Foundation suggests

that granting agencies and foundations fund specific

campus projects that have the potential for success and

could serve as a model for other institutions to use.

Additionally, the report recommends “seed money” funding

for projects that will eventually sustain themselves (Strauss

1996).

Other potential funding mechanisms include a university

budget, internal “loans” repayable with savings from cost

avoidance programs, and a student environmental fee. The

more innovative the design of the OEA, the more marketable

it is; as such, the OEA could easily raise outside funding—

especially from alumni and foundations. Other necessary

resources include personnel, information and data, and an

office. The issue of personnel is addressed later. Initial

sources of information and data already exist (the larger

study from which this paper is extracted and the

environmental audit mentioned above), but an annual report

of the OEA submitted to the TEC, e.g., the “State of the

Tulane Environment”, could continue to chronicle important

information and data. Finally, the OEA has been allocated

office space in the new Environmental Science Building

complex where it will be in close proximity to most of

Tulane’s environmental research and education programs.

The CBR, Green Club, and Environmental Studies Program

can provide necessary office supplies, including a computer,

until full funding is raised.

Leadership

Empower the OEA to make a positive impact on campus.

The Environmental Coordinator should work closely with

various campus entities and constituents to develop and

implement greening initiatives.

Means and Ends

Educate the campus on environmental issues. This

education could be via the following possibilities: large- and

small-scale seminars and programs for students, staff,

faculty, and administrators; continued research into and

implementation of greening initiatives; a comprehensive

measurement/reporting system; the development of an

environmental management plan; and classroom and

curriculum initiatives. The TEC should initially prioritize

projects, and, after the first year, the TEC could approve

annual plans and review past performance. The “ends”

should be outlined in general and specific policies. The

Environmental Coordinator might also teach environmental

classes such as “The Campus and the Biosphere” or

“Ecological Design.”
10

THE OFFICE OF ENVIRONMENTAL AFFAIRS

Leadership

The OEA will house the leadership that will make

environmental change at Tulane: the Environmental

Coordinator. As the director of the OEA, the Environmental

Coordinator should report to the TEC. Dr. John McLachlan

and the CBR would essentially provide a “home” and some

day-to-day operational oversight for the OEA, whereas the

TEC would provide the approval and guidance for longrange operations. Dr. McLachlan might also chair the TEC.

Such an establishment is necessary because of the access

to the varied power and resources of TEC members in

addition to the valuable experience with successful

environmental change initiatives of the CBR and its director.

The TEC would involve the people who guide the university

in its daily and long-range operations and would ensure that

environmental concerns are heard. The TEC could appoint a

working group (with ample student involvement) to

cooperate with the OEA throughout the year on projects and

programs.

The OEA should be “bootstrapped” to each division and tier

of the university: research, education, and operations; and

students, staff, faculty, and administrators. Bootstrapping

involves creating official and unofficial connections that

prevent atrophy or abolishment and that foster collaboration

and cooperation between all areas of the university. Such

connections include research programs with the CBR;

educational and service programs with the Green Club, the

ENST, and various deans; and operational programs (the

ones that will receive much of the focus) with the appropriate

vice presidents and facilities administrators. Many other

connections are possible with Janitorial Services, Student

Programs, Orientation, Admissions, Housing and Residence

Life, Athletics, and campus institutes. These connections

bootstrap the OEA to the core of the university and provide

ways to affect change.

Having an environmental coordinator—the leader—is

absolutely critical to the institutional environmental change

movement. The leader should be a full-time employee with

appropriate experience and degrees; the leader cannot be a

student, although students are the second key to success.

Students

Students from the Green Club, ENST, student organizations,

and the general campus population are pivotal to the

feasibility and success of the OEA. Not only could students

carry out office duties in the OEA, they could also participate

in and benefit from the myriad programs. To maintain their

involvement, ENST and OEA fundraising endeavors could

provide work-study funds for student workers, scholarships

for leadership and academic excellence, and research

assistanceships for student projects. Such opportunities

would also be excellent recruiting tools.

As “customers,” students are effective advocates for change;

they could advocate and stand up for issues in student

milieus by representing the OEA on various campus

committees. Through the OEA, students would have an

organized outlet for environmental activism, volunteerism,

and research opportunities as soon as they arrive on

campus. They would provide a constant source of

enthusiasm and ideas for the program, continually clarifying

the raison d’etre of the OEA. Students would be involved in

an active learning and service community, and they could

gain valuable leadership and job skills by, for example,

taking part in efforts to educate campus denizens through

various programs such as greening seminars, the Internet,

and publications.

As well as contributing to the success of the programs of the

OEA, the students will also be active participants in their

own education. In addition to classes, other educational

venues include service learning in the community and

campus environmental research, effectively using the

campus as a laboratory for environmental problem solving

and for learning how to make positive environmental

change.

Programs of the OEA could also help create connections for

students, especially between students and place, i.e.,

Tulane and New Orleans. The connections they make at

Tulane through the OEA—with outside agencies, community

members, with professors and, most importantly, with each

other— would ensure the lasting success of the OEA

because of the broad and dedicated alumni support

network that could develop. And the innovative programs of

the OEA and ENST could certainly attract talented new

students.

The OEA will depend integrally on students; it will also

empower, support, and educate them. The relationship will

be one of symbiotic, collective leadership, and learning.

Campus sustainability programs are an extraordinary boon

for the students, the entire university community, and,

ultimately, modern civilization. The students will carry their

lessons and skills with them, disseminating environmental

sustainability wherever they live.

Programs

Potential programs of the OEA range from large-scale

projects (conferences with national or international

organizations) to smaller scale projects (office recycling

education in a department; they would encompass the four

divisions and four tiers of the institution, the areas of

Tulane’s strategic interest, and all appropriate environmental

parameters. Through the TEC, presidential invitations could

be sent to key faculty and administrators to strongly

encourage them to participate. In doing so, the OEA would

educate campus decision-makers and crystallize their

involvement with campus stewardship programs, all of which

would strive for ecological literacy.

The OEA would not necessarily run all the programs, but it

would help coordinate efforts, provide information and

experience, and advocate for new programs. Students are

an integral part of the programming function of the OEA, and

they comprise the crucial links between the OEA and the

myriad departments, programs, and organizations on

campus and in the community. The successful projects of the

OEA should be chronicled in campus, local, and national

media. Projects would likely begin focused on campus; once

the OEA builds momentum and accomplishes some campus
11

tasks, programming could move into the local community.

The program possibilities are virtually endless.11

CONCLUSIONS

In the early fall of 1998, it was estimated that, with

fundraising throughout the year and with hiring in the spring,

the OEA could be in place by the summer in order to prepare

for the next academic year. In 1998 when an earlier version

of this study was circulated to raise support and funding,

eight deans, the CBR director, the vice president for

administration and planning, and the provost all said that

they were in support of the initiative in principle. Funding was

solicited from these senior administrators via a one-page

proposal abstracted from the “Blueprint”. It cited the research

findings (presented earlier) and recent events (outlined later)

as sources of campus support and asked for their financial

contribution. Follow-up meetings ensued with most of them.

Eventually, all were in support but were hesitant to commit

resources before they knew the opinions of the president.

After convincing them of presidential support, each donated

amounts ranging from < $1000 to over $5,000. Then the

hiring process began, and the OEA was established.

In 1998–99, Tulane was in a time of profound change—a

presidential transition. Tulane’s new president, Dr. Scott

Cowen, saw that year as a “renaissance of thought and

action” to redesign Tulane for the future. That time of

strategic planning was an opportune time for

institutionalizing the greening process, and the grassroots

advocacy pulled out all stops in order to convince President

Cowen that the OEA was a good and worthwhile venture. He

was not immediately convinced of the validity of the project,

but after 1 year of advocacy, he gave his verbal support. At

least seven actions were fundamental to the advocacy of

1998–99. All were student initiated and led, thus exhibiting

the unique role and power of students in implementing

change. The president was their ultimate target, but the

advocacy helped crystallize the involvement of the

administration and the campus in general.

First was the establishment of the Associated Student

Body’s Committee to Green Tulane, which created a

proposal asking the administration, and the president

specifically, to implement steps that would make the

university more environmentally responsible, including

establishment of the OEA. By the spring, the committee

presented its recommendations to the student body

assembly, which passed the resolution; a subsequent

campus newspaper article on the resolution included a

quote from President Cowen supporting the

recommendations in principle. Second was a year-long

series of letters to the editor regarding campus

environmental initiatives, all of which were in response to

contemporary campus news issues, such as parking and

security lighting, which were not originally presented as

environmental in nature. The third item involved the

collective results of numerous programs by the Green Club,

including public events on Earth Day and America Recycles

Day as well as numerous smaller events that garnered

campus media attention. Fourth was a continual series of

meetings on the proposal with the president and other

administrators. Fifth was the attention provided to Tulane

when the present author, then a graduating senior, was

selected for a national scholarship, as well as top campus

honors, for work greening Tulane; those events involved

local television as well as local and national newspaper

coverage.

The penultimate item was of great importance to the

eventual success of the advocacy. Using the Greening the

Campus study and some of the core texts that informed it

(Creighton 1998, Keniry 1995), the present author and Dr.

Charles Reith, a visiting professor of business and

environmental studies, teamed up to create a class entitled

“Ecological Design.” The class performed an in-depth audit

of Gibson Hall, Tulane’s main administration building that

houses the offices of the president, provost, senior vice

presidents, and other upper administration. The results of

the audit were presented in poster form at a campus-wide

Earth Day celebration in Alcee Fortier Hall, one of the two

new environmental science buildings on campus. President

Cowen, many administrators, and hundreds from the

campus community were present at the event. A report with

suggestions for specific improvements, one of which

involved the establishment of the OEA, was prepared and

submitted to President Cowen; the campus faculty-staff

newsletter featured the Gibson Hall audit, and the report and

posters were placed in Gibson Hall itself as well as on the

Internet.12 In addition, earlier in the semester members of

the class attended and presented at a national conference

on ecological campus design. The eight Tulane delegates—

two staff, one faculty, three students, and the two

professors—eventually drafted a letter to the administration

supporting efforts promoted at the conference, and they

formed a working group that continued to meet and follow up

on ecological design issues. The class also sponsored two

campus-wide events: a public lecture by renowned campus

environmental activist David Orr of Oberlin College, which

was well attended (especially by administrators) and was

pivotal to the campus greening dialogue; and a round-table

discussion on the merits of the International Standards

Organization’s environmental management scheme, ISO

14001.

The ISO issue became the final and pivotal item in the

advocacy effort. President Cowen, a professor of business,

understood the language of ISO 14001 and eventually lent

his support to a concerted effort to decide if it was best for

Tulane. A 1-day conference on the feasibility of ISO at Tulane

was held in May of 1999; it involved many faculty and senior

administrators, and even included an important appearance

by President Cowen. The proposal for the OEA (the

“Blueprint”) now included language that called for pursuing

ISO certification, which Tulane saw as a program that could

eventually go beyond campus boundaries and connect with

11 For examples of potential programs, see Greening the Campus

(footnote 1), especially Appendix F (also Table 4 of the Executive

Summary).

12 The class report and syllabus are at www.tulane.edu/~greenclb/

enst481/ .
12

local institutions that have environmental impacts

(especially industry).

With presidential support and 2 years of funding procured,

a national job search began in early June. In early August, a

committee hired Elizabeth Davey, who had experience in

campus greening initiatives and the added bonus of having

a Ph.D. in English, providing her with a familiarity with the

academic environment and a certain legitimacy among the

faculty. Unfortunately, the group who formed the search

committee and who should have eventually become the

Tulane Environmental Committee, did not form the TEC, and

thus starting institutional change and campus greening

moved more slowly than it would have if the TEC were

initially in place. Whereas the TEC has yet to be formed, the

ISO initiative may result in its creation because it is an

important element for ISO certification. Its absence has been

noticeable.

It was a controversial decision to place the OEA in the CBR

because of the center’s isolation from many aspects of the

university (especially the operations). However, the most

appropriate home for the OEA, the office of the vice

president for administration and planning, agreed to

consider relocation of the OEA after it becomes established

because placing it in the CBR for administrative start-up

purposes seemed the best decision despite the political

ramifications. The office space in one of the two new

environmental studies buildings on campus guaranteed a

close proximity to other environmental programs such as the

Environmental Studies, the CBR, and the Green Club,

whose Student Environmental Center is located there. After

a year of operation, the future placement of the OEA is still

uncertain.

All the elements of model for institutional change are evident

in the events that unfolded at Tulane. Presenting the model

to the administration also legitimized and supported

proposed decisions since they saw the logic, data, and

support of the literature (and academics resonate with

academically supported arguments). The advocacy stage

involved the collective efforts of many to convince the

administration to go forward with the proposal; the TEC as a

formal entity was not the instrumental advocacy group as in

the “Blueprint.” Resources were pooled collectively, and

plans are already in place for alumni fundraising and grant

writing. Policy exhibited the dynamism provided for in the

model. Only informal policy (the early versions of the

“Blueprint” and commitments to funding support, for

example) was in place before the hiring of Dr. Davey.

Eventually, formal policies will be established as in Tulane’s

Strategic Plan in which, as a result of Davey’s work and the

advocacy that led to her appointment, President Cowen has

recently included environmental issues. The OEA is in place,

and its leader, Dr. Davey, is working closely with the various

groups, departments, and schools and is developing a

variety of means to educate the campus on environmental

issues with the eventual ends of environmental

responsibility. Although much planning remains to be done,

she has already started campus greening projects such as:

instituting a grant program for student campus environmental

research and education projects, creating with alumni the

Tulane Environmental Network, improving recycling and

composting efforts, creating an “ecological design” chapter

of the campus master plan, developing a preorientation

program, and developing an environmental Web page.13

The model has provided a helpful framework for

understanding and activating institutional change. The

limitation of the “Blueprint” (and of the model) has been

recognized: it limits to one person the responsibility for

making Tulane an environmentally responsible institution,

rather than sharing the responsibility and initiative across

schools and departments. Some of that limitation could have

been alleviated if the TEC had been successfully established

before or at the time of Davey’s hiring. The second element

to remove some of that limitation, however, is being

actualized. With Davey’s leadership, students are playing key

roles in greening Tulane; they are being supported with

work-study funding, semester research grants, and

extracurricular activities. In fact, it cannot be stressed

enough that the students—from the Green Club, the

Environmental Studies Program, the Associated Student

Body and the general campus population—were absolutely

central to the success of the entire initiative to institutionalize

campus greening. Moreover, they are key to maintaining the

effort, and, after their campus tenure, these students and

their peers will take sustainability beyond the campus to

create a more sustainable world.

ACKNOWLEDGMENTS

The Tulane Environmental Studies Program, the Center for

Bioenvironmental Research at Tulane and Xavier

Universities, and the Ecological Society of America provided

research funding and administrative support. I would like to

thank Timmons Roberts, Michael Zimmerman, Charles

Reith, John McLachlan, Teresa Soufas, Yvette Jones, and

Christine Murphey for their support and thoughtful

suggestions. Heartfelt thanks also go to Julian Keniry and

David Orr for inspiration and stimulating conversations.

REFERENCES

Ackerman, F. 1997. Why do we recycle? Markets, values, and public

policy. Washington, DC: Island Press. 210 p.

Alinsky, S.D. 1971. Of means and ends. In: Rules for radicals. New

York: Vintage Books: 24–27.

Altbach, P.G. 1974. University reform. Cambridge, MA: Schenkman

Publishing. 211 p.

Berry, J.K.; Gordon, J.C., eds. 1993. Environmental leadership:

developing effective skills and styles. Washington, DC: Island

Press. 286 p.

Blueprint for a green campus: the campus earth summit initiatives

for higher education. 1995. The Heinz Family Foundation. 63 p.

Bowers, C.A. 1997. The culture of denial: why the environmental

movement needs a strategy for reforming universities and public

schools. Albany, NY: State University of New York Press. 277 p.

Brown, J.S.; Duguid, P. 1996. Universities in the digital age. Change.

July/ August. 28/4: 10–19.

13 See www.tulane.edu/~eaffairs/ .
13

Cerych, L.; Sabatier, P., eds. 1986. Great expectations and mixed

performance: the implementation of higher education reforms in

Europe. Trentham, Stoke-on-Trent, United Kingdom: Trentham

Books Ltd. 276 p.

Creighton, S.H. 1998. Greening the ivory tower: improving the

environmental track record of universities, colleges, and other

institutions. Cambridge, MA: MIT Press. 337 p.

De Young, R. 1986. Encouraging environmentally appropriate

behavior: the role of intrinsic motivation. Journal of Environmental Systems. 15: 281–292.

Dolence, M.G.; Norris, D.M. 1995. Transforming higher education: a

vision for learning in the 21st century. Ann Arbor, MI: Society for

College and University Planning. 100 p.

Dominick, C.A. 1990. Revising the institutional mission. In: Steeples,

D.W., ed. Managing change in higher education. San Francisco:

Jossey-Bass: 29–36.

Eagan, D.J.; Orr, D.W., eds. 1992. The campus and environmental

responsibility. San Francisco: Jossey-Bass. 133 p.

Eagan, D.; Keniry, J. 1998. Green investment, green return: how

practical conservation projects save millions on America’s

campuses. Washington, DC: National Wildlife Federation. 77 p.

Fantini, M.D. 1981. On effecting change in educational bureaucracies. Education and Urban Society. 13: 389–416.

Farmer, D.W. 1990. Strategies for change. In: Steeples, D.W., ed.

Managing change in higher education. San Francisco: JosseyBass: 7–18.

Gittell, M. 1981. Introduction. Education and Urban Society. 13:

389–398.

Hamburg, S.P.; Ask, S.I. 1992. The environmental ombudsman at

the University of Kansas. In: Eagan, D.J.; Orr, D.W., eds. The

campus and environmental responsibility. San Francisco:

Jossey-Bass: 55–63.

Keniry, J. 1995. Ecodemia: campus environmental stewardship at

the turn of the 21st century. Washington, DC: National Wildlife

Federation. 222 p.

Lane, J.E. 1990. Institutional reform: a public policy perspective.

Brookfield, VT: Dartmouth Publishing Company Limited. 259 p.

MacTaggart, T.J.; Christ, C.L. [and others]. 1996. Restructuring

higher education: what works and what doesn’t in reorganizing

governing systems. San Francisco: Jossey-Bass Publishers.

260 p.

Mansfield, W.H. 1998. Taking the university to task. World Watch.

May/June: 24–30.

Orr, D.W. 1990. The liberal arts, the campus, and the biosphere.

Harvard Educational Review. 60/2: 205–216.

Orr, D.W. 1992. Ecological literacy: education and the transition to a

postmodern world. Albany, NY: State University of New York

Press. 210 p.

Orr, D.W. 1994. Earth in mind: on education, environment, and the

human prospect. Washington, DC: Island Press. 213 p.

Orr, D.W. 1995. Educating for the environment: higher education’s

challenge of the next century. Change. 27/3: 43–46.

Orr, D.W. 1996. Reinventing higher education. In: Collett, J.;

Karakashian, S., eds. Greening the college curriculum: a guide

to environmental teaching in the liberal arts. Washington, DC:

Island Press: 8–23.

Rainsford, G.N. 1990. The demographic imperative: changing to

serve America’s expanding minority population. In: Steeples,

D.W., ed. Managing change in higher education. San Francisco:

Jossey-Bass: 91–100.

Smith, A.; The Student Environmental Action Coalition. 1993.

Campus ecology: a guide to assessing environmental quality

and creating strategies for change. Los Angeles: Living Planet

Press. 129 p.

Steeples, D.W. 1990. Concluding observations. In: Steeples, D.W.,

ed. Managing change in higher education. San Francisco:

Jossey-Bass: 101–107.

Strauss, B.H. 1996. The class of 2000 report: environmental

education, practices and activism on campus. New York: Nathan

Cummings Foundation. 125 p.

Strecker, M. 1998. Tulane makes a billion dollar difference. New

Orleans: Tulane University Public Relations. [December 3 press

release].

Tulane University. 1997. Undergraduate catalog. New Orleans:

Tulane University. 4 p.

Williams, E. 1991. College students and recycling: their attitudes

and behaviors. Journal of College Student Development. 32:

86–88.

Wood, R.J. 1990. Changing the educational program. In: Steeples,

D.W., ed. Managing change in higher education. San Francisco:

Jossey-Bass: 51–58.
14

INTRODUCTION

The University of Mississippi, along with the American

Universities International Program, developed an intense

study-abroad program in Belize since 1995. This

collaboration has seen outstanding success during the past

5 years and involved many university professors, students,

professionals, and other officials from both the United States

and Belize.

PROGRAM OVERVIEW

The Belize field-study program is a unique multicultural

program where participants live and work together with other

students from across the United States and other parts of

the world. They travel throughout Belize with supervisors,

instructors, guides, and local staff to various sites, which

provide for affordable living opportunities to study within the

natural settings and cultures of the local people. Students

are given proper orientation in compliance procedures and

regulations to ensure protection of ecosystems being

studied while maintaining strict North American safety

standards.

The actual classrooms and study sites include spectacular

and pristine ecosystems, which range from the largest

barrier reef in the Western World, lush rain forests,

mangrove swamps, high forested mountains and plateaus,

freshwater lagoons, scenic and wild rivers, extensive

underground caves, Mayan ruins, and much more. The

student is given a truly interactive and hands-on experience

that can definitely put the fun and excitement back into

science education in a most pragmatic and affordable

manner. Such an opportunity can take learning well beyond

the greatest expectations, broaden one’s scope, and

perhaps change one’s overall life perspective in a most

positive manner. Consequently, almost daily, students are in

awe of what they experience and learn.

The current success of the Belize field-study program

demonstrates the tremendous opportunities for science

education and research that American universities can

utilize in collaboration with the many and varied Belizean

institutions. Reliable and dependable cooperation with local

institutions can be readily established that will provide good

accommodations, warm atmosphere, fine food, and a

complex mix of pristine subtropical ecosystems—all at a

super value and within strict safety standards and

regulations.

COUNTRY OVERVIEW

Belize, formerly known as British Honduras, gained

independence in 1981 and remained a parliamentary

democracy and member of the British Commonwealth. It

continues to be a peaceful country, well insulated from the

problem areas of Central America.

Belize is on the eastern coast of Central America on the

Caribbean Sea. It is located between 15o

53' and 18o30' N.

and is bounded on the north by Mexico (the State of

Quintana Roo and in the extreme northwest, Compeche) to

the west and south by Guatemala (the department of Peten

and in the extreme south Izabal) and to the east by the

Caribbean Sea. In form, the country is roughly rectangular,

extending 280 km from north to south and 109 km from east

to west. The maximum extension is 180 km, including the

territorial sea. Total land area, including some 1,200 cays, is

22,960 km2, which is divided into 6 districts, 9 municipalities,

and more than 240 villages. The territorial sea adds to a total

area of 46,620 km2.

In spite of its small size, Belize is composed of a diversity of

landscapes. Inland, the Mayan Mountain/Mountain Pine

Ridge Massif is the dominant physical feature and rises to

1,124 m at its highest point. It is surrounded by rugged karst

limestone hills. Beyond that, most of the north and the entire

coastal area, including the south, consists of low-lying

plains. Nine land regions, each comprising a particular

combination of topography, soils, and vegetation, and, thus,

a distinctive landscape, have been distinguished. The

THE COLLABORATION BETWEEN THE UNIVERSITY OF MISSISSIPPI

AND BELIZE ENABLES OPPORTUNITIES FOR SCIENCE EDUCATION

Richard A. Belisle1

1 Vice President, Beaus, Ltd., Box 300, Belmopan, Belize, Central America.

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability of

wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.

Abstract—The University of Mississippi and the American Universities International Program (AUIP) enjoy vast educational

opportunities in Belize. Bounded by Mexico on the north, Guatemala on the west and south, and the Caribbean Sea on the

east, Belize’s 22,960 km2 of topography range from sea level to 3,688 ft. This variation in altitude and the tropical climate

support a great diversity of habitats. The relative underdevelopment and pristine ecosystems of rain forest, pine savannahs,

and mangrove swamps are complemented by the longest barrier reef in the Western Hemisphere. Since 1995, the primary

purpose of the AUIP study program in Belize has been to provide experiential learning courses for students, professors, and

professionals from the United States and Belize. The main emphasis has been on marine biology, ecology, natural resource

management, anthropology, ecotourism, and land use.
15

upland of the Mayan Mountain land system constitutes the

dominant topographical feature. The geology of this area

comprises Carboniferous and Permian metasediments

(quartzites, shales, slates) with granite intrusions. Most of

this land system is characterized by steep slopes and

shallow soils, which are leached, acidic, infertile, and

fragile.

There are five upland regions, which are characterized by

hilly terrain made up of a thick section of indurated

Cretaceous limestone that has eroded into an extremely

rugged scenery known as karst. These regions are: the

central foothills, located on the northern flank of the Mayan

Mountains Massif; the western foothills, located in the

western and midwestern areas of the country that include

the most extensive cave system in Central America; the

eastern foothills, overlooking the coastal plains and hosting

the oldest rocks in Central America; the southern foothills,

located in the northwestern parts of the southernmost

district; and the Bravo Hills, located in the northwest area of

Belize.

The other three land systems are comprised of the coastal

plains. The southern coastal plain lies in the most southerly

district and is further divided into the Machaca (marginal

agricultural suitability) and Temash (poorly drained acid and

infertile) plains; and the northern and central coastal plains,

both underlain by Tertiary limestone, marl, and calcareous

clays.

The preliminary classification scheme for the coastal zone,

as developed under the current Protected Areas System

Plan, further distinguishes four regions that reflect

differences in sediments, bathymetry, topology, and

hydrology, and represent marine equivalents of the

terrestrial land systems. These four coastal regions, running

from north to south, are the Ambergris region, the Belize

region, the Stann Creek region, and the Punta Gorda

region.

The climate in Belize is subtropical with rainfall varying from

< 1300 mm per year in the north (with a 4-month dry

season) to over 4500 mm in the south (with a shorter dry

season). This remarkable variation in rainfall over a small

area (280 km in length) is incredible but true.

The climate, along with the diversity of geology, topography,

and edaphic factors, maintains pristine ecosystems of rain

forest, pine Savannahs, jungle mountains, rivers, deep

caves, lagoons, estuaries, huge waterfalls and more.

Likewise, the natural vegetation shows astonishing variation

and diversity with some 49 distinct types already recognized.

Together, the terrestrial ecosystems of Belize display as

much habitat diversity at low altitude as any other

Mesoamerican country. The exceptional inland ecosystems

of Belize are home to over 500 species of birds, 107 species

of reptiles, more than 120 species of mammals, 26

amphibian species, and over 700 plant species.

Seven principal habitats make up the Belizean coastal zone:

coral reefs, sea grass beds, mangrove forests, littoral forests,

watersheds, wetlands, and estuaries. The shallow crystal

clear waters of the coastal zone of Belize are host to the

largest barrier reef of the Northern Hemisphere. This is one

of the most spectacular, complex, and biologically diverse

ecosystems in the world. The barrier reef complex runs from

north to south and extends some 230 km along the coastline

from the border with Mexico to the threshold of the border

with Guatemala. Three offshore coral atolls occur east of the

barrier reef and are associated with multiple sand and

mangrove islands enclosing a lagoon with sparse coral

formation. The southernmost atoll, however, has frequent

and well-developed structures.

Some 65 coral reef species have been identified with

suggestions that as many as 113 coral-associated species

may occur. Indications are that apart from corals, as many

as 343 species of other invertebrates may occur. In

December 1996, seven locations on the barrier reef were

designated World Heritage Sites under the World Heritage

Convention; and in 1998, one inland wetland site was

designated under the RAMSAR Convention.

Over 4,000 years ago, the Mayan Civilization took root in

what is now Belize. Some 2 million Mayas may have lived

there during its height. They built sophisticated cities and

temples, and the record of their lives in over 900

archeological finds has survived the years. Although many

temples have been excavated from the earth, many more, it

is certain, remain hidden in the dense jungle. The reign of

the Mayas has vanished, but direct blood ancestors still live

in Belize in harmony with an unlikely mix of newcomers.

They include descendants of English settlers, European

pirates, African slaves, Atlantic Coast Indians, East Indians,

Mennonites, Chinese, Lebanese, and some veterans of the

U.S. Civil War. More recently, an influx of Central American

immigrants has increased the population of Belize to some

230,000 strong.

Although multilingualism prevails in Belize, communication is

not a problem because almost everyone speaks English.

However, the different ethnic groups maintain their cultural

traditions and dialects, thereby adding to the natural charms

of Belize.

Getting to Belize from the United States is facilitated by

direct flights out of the cities of Houston, New Orleans,

Dallas, and Miami, each with a short duration of 2 hours.

Central standard time is observed year round in Belize, and,

similarly, the currency exchange is stable at $2.00 BZE to

$1.00 U.S. Hotels, restaurants, and stores accept U.S. and

Belizean currencies interchangeably.

Belize is still an outpost on the edge of the world, where all

things wild live undisturbed. If the country is famous for

anything, it would be for its genuinely friendly people,

touring, diving along its outer cays, and ecotourism inland.

The richness of the traditional medicine that preserves the

value of many species of plants and animals in the

unspoiled ecosystems is likewise appealing to tourists,

students, professors, and researchers alike. Furthermore,

the availability of modern and highly efficient

telecommunications and Internet facilities throughout Belize

adds irresistible appeal to the high-tech generation of today

and provides exceptional delight to both students and

professors.
16

The primary purpose of the current University of Mississippi

collaborative program in Belize is to provide an experiential

learning adventure whereby undergraduate students, along

with their professors, can develop an understanding of living,

working, and learning in a developing country where pristine

ecosystems prevail. It is designed to entail field study for

persons with backgrounds and interests in natural

resources, biology, geology, ecology, tourism, and

anthropology. It is for anyone who enjoys working and

learning in remote environments and cares about further

conservation of global resources, particularly in a diverse

and challenging natural situation. Special emphasis is

placed on natural and social sciences and how land use and

ecological decisions are developed within existing

socioeconomic constraints.

Some recent direct spin-off activities include: baseline

environmental studies, evaluation and monitoring of

environmental factors, natural product research, research

and development of bioremediation systems and prototypes,

and current and tide data generation within a U.S. Navy

hydrographic project. Also, possibilities exist through

collaboration with the Mississippi Research Consortium in

the completion of a diagnostic for the World Bank’s Regional

Mesoamerican Reef Project. Subsequently, several new

avenues for additional marine biology research activities and

operations have evolved. Clearly, Belize is now being

recognized as a country that affords ideal hands-on

situations for field studies that can certainly provide

tremendous opportunities for science education within a

regional and global context.
17

INTRODUCTION

Constructed wetlands are a feasible solution to treat

domestic wastewater in small rural communities. These

systems are especially valuable for onsite wastewater

treatment in developing countries because they are low tech,

and the costs of construction and operation are low (Denny

1997).

The gains in vegetative biomass in constructed wetlands

can provide economic returns to communities when

harvested. These can be realized through biogas production,

animal feed, fiber for papermaking, and compost (Lakshman

1987). There have been no reports of the use of ornamental

plants, such as commercial flowers, in constructed treatment

wetlands even though these plants can be harvested and

sold.

Most of the research on constructed wetlands has been

conducted in northern countries. Therefore, the plants most

studied for treatment purposes are cattail (Typha spp.),

bulrush (Scirpus lacustris), and reeds (Phragmatis australis)

as these plants are native to northern countries (Denny

1997). It is known that ornamental plants like canna lily

(Canna flaccida), calla lily (Zantedeschia aethiopica),

elephant ear (Colocasia esculenta), ginger lily (Hedychium

coronarium), and yellow iris (Iris pseudacorus) can be used

in rock/plant filters to treat septic tank effluents (Wolverton

1990). However, there is not enough information about the

efficacy of these “warm weather” ornamental plants for use

in treatment wetlands. Because most developing countries

are located in tropical and subtropical areas, the use of

ornamental plants for treatment in wetlands should be

explored. Floriculture opportunities would also provide

economic benefits to the communities in addition to the

benefit of the wastewater treatment.

This study is focused on the use of calla lily in subsurface

flow wetlands to treat domestic wastewater.

OBJECTIVES

The objectives of this study are to set up a lab-scale

subsurface flow wetland to study the feasibility of using the

calla lily to treat simulated wastewater and to assess the

efficacy of the system in removing nitrogen from the

wastewater.

METHODS

A lab-scale subsurface flow wetland was installed in a

greenhouse at Trent University. The system consists of six

cells of 38 cm by 240 cm by 30 cm (width by length by

depth) filled with crushed rock of 3 to 5 cm in diameter. Two

cells were unplanted and four were planted with calla lilies.

The cells were provided with the same simulated domestic

wastewater, stored in an elevated tank, at the same flow. The

output water was collected and pumped back into the tank.

Each cell was planted with 16 plants at a distance of 15 cm

between them.

The plants were 8 cm tall when planted. The system was

maintained at a temperature range of 14 to 20 oC and 16

hours of light per day. After plantation, nutrients were added

to the water in the same composition as is used for

hydroponics to allow the plants to root and grow. After 120

days, once the plants were well established, lawn fertilizer

and tannic acid were added to increase the nitrogen and

carbon levels to those typically found in domestic

wastewater (table 1). The flow was adjusted to operate the

system at 1 day of hydraulic retention time.

Ion-selective electrodes were used to measure N-NH3

, NNO3

, N-NO2; phosphate, sulfate, and chloride were

measured by ion chromatography. Total organic carbon

(TOC) was measured using a TOC analyzer, Shimadzu

model TOC-5000; pH and Eh were also measured. The

monitoring of the water was performed according to

techniques described in the Standard Methods (American

FEASIBILITY OF USING ORNAMENTAL PLANTS IN SUBSURFACE FLOW

WETLANDS FOR DOMESTIC WASTEWATER TREATMENT

Marco A. Belmont and Chris D. Metcalfe1

1 Belmont, Graduate Student, and Metcalfe, Professor, Environmental Resources Studies, Trent University, Ontario, Canada.

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability of

wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.

Abstract—Constructed wetlands are possible low-cost solutions for treating domestic and industrial wastewater in

developing countries such as Mexico. However, treatment of wastewater is not a priority in most developing countries unless

communities can derive economic benefit from the water resources that are created by the treatment process. As part of

our studies directed at improving the quality of water in the Rio Texcoco in central Mexico, we are determining the feasibility

of using ornamental flowers for treatment of domestic wastewater. In a laboratory-scale study, we determined that subsurface flow wetlands planted with calla lilies could reduce levels of ammonia and nitrate in simulated domestic sewage.

Results are presented on the optimal conditions for treatment of domestic wastewater in these systems. Floriculture

activities in constructed wetlands could provide the economic benefits necessary to encourage communities in developing

countries to maintain wastewater treatment systems.
18

Public Health Association and others 1998). The parameters

were measured in input and output water of the cells.

The simulated wastewater was prepared to meet the usual

characteristics of sewage as shown in table 1 (Metcalf and

Eddy 1991).

RESULTS AND DISCUSSION

The experiment shows that the calla lily can grow very well

on crushed rock beds at warm temperatures. The growth of

the plants was faster than expected, and the plants started

flowering after 6 weeks of plantation.

Ammonia and nitrate measurements showed that the

wetland removed ammonia (fig. 1), but nitrate reduction

occurred only at the beginning of the experiment before the

addition of organic nitrogen (fig. 2). In fact, once fertilizer was

added, the nitrate concentration was higher in the output

than in the input water. This probably means that the

production of nitrate by the conversion of organic nitrogen

into ammonia followed by conversion of the ammonia to

nitrate exceeded the nitrate removal capability of the system.

In other words, the concentration of nitrate increases

because of the lower capability of the system for

denitrification than for mineralization and nitrification. It is

necessary to quantify organic nitrogen to support this

hypothesis.

Nitrite was found in very low concentrations compared with

nitrate and ammonia; therefore, nitrate and ammonia are

more descriptive of the nitrogen transformation in the

artificial wetland.

Measurements of pH showed that the output water from the

wetland had a quite constant pH, slightly higher than 7, even

when the input water pH changed with time for almost 2

units (fig. 3). The redox potential (Eh) in the input and output

water was very similar through the sampling period (fig. 4).

This parameter showed a tendency of decrease with time

throughout the duration of the experiment. This is probably

due to the addition of organic nitrogen, ammonia, and

organic carbon to the input water.

CONCLUSIONS

Calla lily plants can live and grow very well in subsurface

flow beds of crushed rock. In warm places, the calla lily can

be used in subsurface flow wetlands to treat wastewater and

produce cut flowers. This adds a financial benefit to the

environmental benefit of wastewater treatment. Under proper

Table 1—Typical water-quality parameters in domestic

wastewater

Sewage in North America Experimental wetlanda

Parameter medium-strong April-May 2000

 - - - - - - - - - - - - - - - - - Mg/L - - - - - - - - - - - - - - - -

TSS 720 – 1200 6.5 – 14.5

TOC 160 – 290 110 – 155

N-NH3 25 – 50 70 – 180

N-NO3 0 – 0 60 – 170

N-NO2 0 – 0 1.2 – 11.16

pH 6 – 8 6.9 – 8.1

Phosphate 5 – 10 25 – 75

Sulfate 30 – 50 600 – 620

Chloride 50 – 100 43 – 57

a Metcalf and Eddy 1991.

Figure 1—Ammonia concentration in mg/L.

Figure 2—Nitrate concentration in mg/L.

Figure 3—Measured pH.

NH3

pH

Input

NO Cally Lily 3
19

conditions of hydraulic retention time, calla lilies can be used

in subsurface flow wetlands to reduce levels of nitrogen and

to convert ammonia to nitrate.

For the next phase of the experiment, it is necessary to

quantify total organic nitrogen. The artificial wetland

regulates the water pH buffering changes in the input water

pH.

Figure 4—Redox potential (Eh) in mV.

ACKNOWLEDGMENTS

We thank the Canadian International Development Agency

for the financial support received through the INSTRUCT

project and Consejo Nacional de Ciencia y Tecnologia

(CONACyT, Mexico).

REFERENCES

American Public Health Association; American Water Works

Association; Water Environmental Federation. 1998. Standard

methods for the examination of water and wastewater. 20th ed.

Washington, DC.

Denny, P. 1997. Implementation of constructed wetlands in

developing countries. Water Science Technology. 35(5): 27–34.

Lakshman, G. 1987. Ecotechnological opportunities for aquatic

plants—a survey of utilization options. In: Reddy, K.R.; Smith,

W.H., eds. Aquatic plants for water treatment and resource

recovery. Orlando, FL: Magnolia Publishing Inc.: 49–68.

Metcalf and Eddy Inc. 1991. Wastewater engineering. Treatment,

disposal and reuse. 3rd ed. New York: MacGraw Hill Inc.:

109–110.

Wolverton, B.C. 1990. Aquatic plant/microbial filters for treating

septic tank effluent. In: Hammer, D.A., ed. Constructed wetlands

for wastewater treatment. New York: Lewis Publishers: 173–185.

Calla Lily

Eh Input
20

INTRODUCTION

Burgeoning human populations with inherent needs for food

and income continue to drive the settlement and exploitation

of coastal regions. Some 20 percent of the human

population (over 1 billion people) live within 30 km of the

coast (Gommes and others 1998). In tropical developing

countries, approximately 90 percent of fishery landings

come from shallow coastal waters and provide for 40 to 90

percent of national animal protein consumption (Holdgate

1993). In Asia, it has been estimated that one billion people

depend exclusively on fish for their protein requirements

(Anon. 1985). Population densities in excess of 1,000 people

per square kilometer are commonplace in rural areas of

developing countries in Asia, the Pacific, Central America,

and the Caribbean (Lundin and Linden 1993). Such

population densities place adverse pressures on coastal

zones, which are used for settlement, transport, waste

disposal, agriculture, forestry, aquaculture, and fishing. In

many cases, the sustainable levels of exploitation in

fisheries, harvest of mangroves, or the use of the

assimilative capacity of wetlands to deal with sewage has

long been exceeded. Competition for and overuse of

renewable resources have been identified as a major

problem in many regions. There are also mounting conflicts

between different forms of resources development that

reduce the effectiveness of investment and threaten the

sustainability of resource production. For example, logging

and mining activities in upland areas of Southeast Asia have

brought short-term and localized economic benefits to the

communities involved. At the same time, they have caused

major environmental damage and imposed negative

economic impacts that damaged capture fisheries,

aquaculture, and tourism interests in lowland areas (Chou

and others 1991, Hodgeson and Dixon 1988).

We are therefore facing a very serious challenge in most

coastal regions where rapid growth in coastal populations,

rapid urbanization, competition for land and water

resources, and pollution are undermining the potential of

coastal zones to sustain social and economic development

objectives. One such objective is sustainable protein

production through wise fisheries management and the

development of aquaculture. However, throughout the world,

coastal ecosystems believed to play a significant role in

supporting fisheries and aquaculture are being lost or

severely degraded. Coastal wetland forests play such a role

but are not being effectively conserved. For example,

estimates suggest that 70 percent of Thailand’s mangrove

has been converted to industrial sites, agriculture, and

shrimp ponds. A policy to encourage the conversion of

mangrove and other coastal ecosystems into shrimp ponds

was stimulated by the reduction of open-access fishing

areas as a result of the declaration of Exclusive Economic

Zones by countries bordering the Gulf of Thailand. The

policy aimed to relieve unemployment among fishermen

and to increase fisheries production. However, little or no

account was taken of the impact of the loss of mangrove in

terms of decreasing support of capture fisheries or

aquaculture. This lack of attention to such a basic issue

may undermine the benefits derived from aquaculture when

good planning and management could avoid a conflict of

interests between forestry, aquaculture, and capture

fisheries.

With the mounting pressures on coastal areas and

resources, it is increasingly important to strive for the

sustainable use of ecosystems and the renewable natural

resources they generate. However, this takes sophisticated

development planning and careful management of

development activities. In the case of shrimp-pond based

aquaculture development, we are facing two distinct sets of

issues affecting the sustainability of production. The first is

the adverse impact of other forms of development on

aquaculture; the second is the poor siting and management

of the aquaculture units.

REHABILITATION OF COASTAL WETLAND FORESTS DEGRADED

THROUGH THEIR CONVERSION TO SHRIMP FARMS

Peter R. Burbridge and Daniel C. Hellin1

1 Burbridge, Professor, and Hellin, MSc, Department of Marine Sciences and Coastal Management, University of Newcastle, Newcastle upon

Tyne, NE1 7RU, England.

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability of

wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.

Abstract—International demand for shrimp has stimulated large-scale conversion of mangrove and other coastal wetlands

into brackish water aquaculture ponds. Poor site selection, coupled with poor management and over-intensive development

of individual sites, has led to nonsustainable production and often, wholesale abandonment of ponds. This has been

followed by further conversion of wetlands in an attempt to maintain aquaculture production, incomes, and employment.

This has also often proved nonsustainable. The net result is that extensive areas of formerly biologically rich and productive

wetland forest are lying idle. In limited cases, natural regeneration of wetlands is taking place, and there are sporadic

attempts to stimulate regeneration. However, the drive to convert further wetlands is far greater than efforts at rehabilitation.

The development of alternative, sustainable uses of former wetland forests is examined as a means of reducing the

pressures to convert further areas of wetland forest.
21

Aquaculture has a legitimate right of access to and use of

coastal resources as well as a right not to be adversely

affected by the poor planning and management of other

forms of coastal development. These issues have been

addressed in a comprehensive manner in recent reports of

the Working Group on Environmental Interactions of

Mariculture (ICES 1995, 1996, 1997, 1998, 1999). The

second set of issues has been addressed by FAO (Barg

1992) and numerous other technical groups and individual

experts. For example, as early as 1981, Juliano and others

(1981) were able to demonstrate that production from shrimp

ponds in the Philippines could be doubled through marginal

improvements in the management of ponds using existing

knowledge, thus reducing the need to expand the area of

mangrove converted to aquaculture.

Agencies such as the World Bank and FAO are addressing

basic institutional and technical issues concerning the

establishment of coastal management in developing nations.

However, no international organization has yet developed

proactive approaches or practical guidelines for the

integration of aquaculture into coastal planning and

management activities in both developing and developed

nations. Where integration is being attempted, it is based

upon local or regional planning within individual nations

(ICES Working Groups on CZM 1995). As a result,

opportunities for promoting the sustainable growth and

diversification of coastal aquaculture are not being fully

realized, and options for future development are being

foreclosed.

Despite the feasibility of greatly enhancing the sustainable

production of aquaculture through integrated development

planning and sound management, there has been a rapid

and largely unplanned expansion of shrimp aquaculture in

developing nations. As a consequence, there has been

large-scale destruction of mangrove and other wetland

systems with consequent impacts on human activities that

are supported directly, or indirectly, by their ecosystem

functions. A second consequence has been the cessation of

production in extensive areas of aquaculture sites. Land is

frequently lying idle, yet there is continuing pressure to

convert more wetlands in an attempt to maintain

aquaculture production. This trend cannot be sustained, and

it is argued that rehabilitation of disused and unproductive

aquaculture sites presents a positive means of bringing

these lands back into some form of productive and

sustainable use while reducing the pressure to convert

remaining productive wetlands. This presents a major

challenge to all of us involved in forestry, aquaculture,

capture fisheries, and other aspects of human development

in coastal regions.

THE ISSUE OF NONSUSTAINABLE CONVERSION

OF COASTAL WETLANDS TO AQUACULTURE

The actual scale of conversions of mangrove and other

wetlands to aquaculture is very difficult to document. It has

been estimated that over 15 million ha of mangrove had

been cleared. This accounts for approximately half of the

mangrove that had previously existed, and this is thought to

be further decreasing at a rate of between 2 and 8 percent

per year (Kunstadter and others 1985). In the Philippines,

279,000 ha of mangrove were lost between 1951 and 1988,

and conversion to culture ponds accounted for

approximately 50 percent of this loss (Primavera 1995).

In addition to areas converted from mangrove to shrimp

ponds, other areas have been converted from salt flats, salt

marshes, freshwater wetlands, fishponds, rice paddies, or

agricultural lands. Frequently, ponds were built in regions

already degraded by other practices and already in need of

active management.

Aquaculture of marine species has been a well-established

feature of coastal development for more than a thousand

years in parts of Asia. An example of this is the polyculture

of milkfish and shrimp in “Tambaks” (brackish water ponds)

in Indonesia. Most of these systems were extensive in

nature and many integrated mangrove conservation into the

pond management system as both a means of enriching the

food supply in the ponds and of protecting ponds from

coastal storms. The traditional extensive or semi-intensive

pond systems have largely given way to intensive systems,

which require very careful site selection and high levels of

inputs of feed and other materials to balance water

chemistry and maintain the health of the cultured species.

The capital investment, technical skills, and knowledge of

markets required to sustain production in intensive

aquaculture systems often puts them beyond the reach of

most rural communities. Where they are attempted without

the requisite inputs and protection from adverse external

impacts, production cannot be sustained, and they become

disused.

The Scale of Disuse of Aquaculture Sites

Accurate estimates of pond disuse (in both mangrove and

nonmangrove areas) are difficult to obtain because land

tenure records are often unreliable and out of date, and

assessments using remote sensing are hampered by the

inability to discern between productive and disused ponds

(Stevenson and others 1999). Unofficial estimates of pond

disuse have suggested that the percentage of ponds left idle

after a period in production can be as high as 70 percent

(Stevenson 1997). Attempts to quantify the scale of pond

disuse have been marred by the belief that an admission of

pond abandonment is tantamount to an admission of

management failure, and, to date, comprehensive surveys of

disused shrimp ponds have not been undertaken

(Stevenson and others 1999). In practice, ponds are often

converted to other uses; for instance, in Thailand, some

ponds have been sold for housing and industrial

development, converted to salt farms or fish or crab culture

operations, and some shrimp farmers have sold topsoil for

construction projects (Stevenson 1997).

In Thailand, Potaros (1995) stated that 19 900 ha of shrimp

farms in the five provinces of the Inner Gulf of Thailand were

closed in 1990–91. A report produced by the Network of

Aquaculture Centres in Asia-Pacific (NACA) indicates that in

1989 about 62 percent of farms were operating “under

capacity” and another 22 percent of farms were “abandoned”

in Samut Sakhon province (Office of Environmental Policy

and Planning - OEPP 1994). This is supported by Briggs and

Funge-Smith (1994) who reported that an area of 40 000 to

45 000 ha south of Bangkok became derelict after shrimp

production collapsed in 1989–90.
22

Yap (1997) reports that nearly all of the 54 912 ha of shrimp

ponds in the Philippines were abandoned, and another

83 000 ha of brackish water ponds were “idle”. Reports

from NGOs in the Philippines in late 1997 stating that pond

disuse is common in the Philippines have supported this,

although pond operators have frequently returned to

traditional forms of milkfish [Chanos chanos (Forsskal)]

culture after shrimp production has ceased (Stevenson and

others 1999).

The disuse or abandonment of coastal aquaculture sites has

been reported elsewhere but not quantified. Extensive areas

of disused shrimp ponds are thought to exist in Bangladesh,

China, Malaysia, and Colombia (Stevenson 1997) and, more

recently, Mexico.

Causes of Disuse and Abandonment of Ponds

There are numerous reasons for the cessation of production

in shrimp ponds. Examples include:

• poor site selection (reported in Sri Lanka by Jayasinghe

1995);

• flooding due to poor catchment management as well as

from storm surges where the buffer function of mangrove

has been lost due to their removal;

• predation by nontarget species in the ponds, e.g., birds

and other animals;

• poor cohesivity of soils, which causes the pond walls to

collapse;

• acidification of soils and water as a result of the exposure

of potential Acid Sulfate Soils (reported in Vietnam by

Tuan 1996, in Cambodia by Sreng 1996);

• contamination of pond water from agricultural wastes

(noted in Indonesia as a result of a shortage of fresh

water and problems of water quality by Burbridge pers.

obs. 1997); and

• diseases resulting from a lack of hygiene, which can be

rapidly transmitted among ponds through poor water

management, for example, reported in India as a result of

white spot disease (Sammut and Mohan 1996), in Sri

Lanka (Jayasinghe 1995), in the Philippines (Ogburn and

Ogburn 1994), and in Taiwan (Stevenson and others

1999).

Briggs and Funge-Smith (1994) were among the first to

highlight the problem of poor hygiene and diseases in a

report to the British Overseas Development Agency (now

the Department for International Development, DFID).

Hambrey (1996) reported that chronic disease and waterquality problems have caused ‘significant’ pond

abandonment. For instance, disease problems have caused

abandonment in India, (Sammut and Mohan 1996), the

Philippines (Yap 1997), and Thailand (Macintosh 1996).

Poor water quality and poor site selection have caused

production failure in Sri Lanka (Jayasinghe 1995) and

Indonesia, and problems with Acid Sulfate Soils (A.S.S.)

have caused abandonment in Vietnam (Tuan 1996) and

Cambodia (Sreng 1996). These problems often lead to

financial difficulty, causing farmers to either sell or abandon

their farms (Fegan 1996). Ponds may also be abandoned

due to a drop in profits or yields (Flaherty and

Karnjanakesorn 1995), or political intervention, such as the

revoking of leases or license agreements (Stevenson and

others 1999).

Where the development of coastal aquaculture has not

been well planned and managed, it undermines the

potential for coastal zones to sustain economic and social

development. This raises the question of whether it would

be better to rehabilitate these areas and restore the original

ecosystem, or to find a means of modifying the aquaculture

system to allow it to be more productive and sustainable.

There are good arguments for both alternatives. On one

hand, the rehabilitation of the original ecosystem may help

to rejuvenate coastal capture fisheries stocks and the

income of fishermen, improve biodiversity, ecotourism, and

reduce salinization of soils and groundwater, which

adversely affect agriculture and domestic water supplies.

However, it must be realized that rehabilitation will cost

money, will take considerable time, and may not be

welcomed by local people who may see little benefit for

themselves.

On the other hand, developing more productive use of

unsuccessful aquaculture sites could allow the original

developers to achieve a reasonable return on their

investment and could provide opportunities to diversify and

expand local employment. Mixed aquaculture systems may

help improve the food security of rural communities and

reduce organic pollution loads from other forms of

development, including more intensive aquaculture. One

example might be the change from nonprofitable intensive or

semi-intensive shrimp culture to a less intensive polyculture

system or an integrated aquaculture-agriculture forestry

system.

These and other alternatives will vary from place to place

depending on environmental, social, and economic

conditions. At this meeting, it would be useful to discuss how

to develop a system to identify and evaluate opportunities to

put nonproductive and idle aquaculture sites into a more

productive use that helps to meet sustainable development

objectives at both a local and national level. In the following

paragraphs, factors that influence options for rehabilitating

disused aquaculture sites are set out.

OPTIONS FOR THE REHABILITATION OF

DISUSED OR ABANDONED AQUACULTURE

PONDS OR BOTH

The term restoration has been adopted in recent studies to

mean any activity that aims to return a system to a preexisting condition, whether or not this was pristine (sensu

Lewis 1990b). Whereas, the term rehabilitation is used to

denote any activity that aims to convert a degraded system

to a stable alternative use, which is designed to meet a

particular management objective (Stevenson and others

1999). The term rehabilitation is used here to describe a

continuum of management options for altering the state of

the ponds to some alternative condition where human

activities can be sustained. This can include the

reinstatement of a wetland forest ecosystem such as a

mangrove, where uses foreclosed through the conversion of

the wetland can be regained and benefit a range of different

interests. For example, mangrove forests have been restored

to meet commercial purposes such as silviculture (Watson

1928), for restoring fisheries habitat (Aksornkoea 1997,

Lewis 1992), for sustainable multiple community use
23

purposes, or for shoreline protection purposes. None of

these are mutually exclusive.

There are three basic rehabilitation options. The first is to

rehabilitate the pond sites so that they can be put back into

sustainable shrimp production. The second is to rehabilitate

the pond sites so that they can be put to some alternative,

sustainable use. The third option is to restore the

environmental conditions within the pond sites and the

surrounding area, and to re-establish a productive wetland

ecosystem (Stevenson and others 1999). Each of these

options is influenced by the causes of production failure and

the conditions that remain in the pond after production has

ceased (Stevenson 1997). The continuum of rehabilitation

options is illustrated in figure 1.

Factors That Influence the Choice of

Rehabilitation Options

The basic rationale for attempting rehabilitation is to address

factors that alter a wetland or other productive coastal

ecosystem to such an extent that it can no longer selfcorrect or self-renew. Under such conditions, ecosystem

homeostasis has been permanently stopped, and the

normal processes of secondary succession (Clements 1928,

Watson 1928) or natural recovery from damage are inhibited

in some way.

Before any restoration is attempted, the goals should be

determined through an active dialogue with, and effective

participation of, the local stakeholders. The stakeholders

comprise both individuals or interest groups that have

promoted aquaculture, as well as those who may have been

influenced by the conversion of wetlands and other systems.

Those with interests in aquaculture may have invested

scarce capital, as well as their labor and materials, to

develop what they saw as a means of improving their

welfare. In the process of developing the aquaculture sites,

they may have attained user rights or title to the land. They

will want to see some productive outcome from restoration

efforts that will help them achieve sustainable return on their

investment. On the other hand, people who did not benefit

from the aquaculture development may have suffered a

loss of access to common property resources such as the

crabs that breed and thrive in mangrove forests or fish and

crustaceans that depend upon wetlands for part of their life

cycle. Balancing the needs and aspirations of the different

stakeholders and achieving broad agreement on the goals

are critical to the sustainability of rehabilitation initiatives.

The importance of this involvement of stakeholders is often

overlooked, and, without local support, restoration and

rehabilitation programs have little chance for long-term

success (Primavera and Agbayani 1996).

It may be possible to restore the functionality of a system

even though factors such as soil type and condition may

have been altered with consequent effects on the flora and

fauna. If the goal is to return an area to pristine condition,

then the likelihood of failure is high and must be regarded as

unrealistic (Stevenson and others 1999). The restoration of

certain ecosystem traits and the replication of some

functionality stand more chance of success (Lewis and

others 1995).

The reference conditions for restoring the functions of a

system also need to be examined carefully because “preexisting conditions” may not have been pristine due to factors

outside those associated with the conversion to aquaculture.

Potaros (1995) reports that in Thailand, “the mangrove area

that has been used for shrimp farms is often previously

degraded forest so it is very difficult to assess the economic

damage related directly to shrimp farming”. In such cases, it

is not sensible to return the area to a previously degraded

condition. The goal may be to achieve a level of functionality

that will allow productive alternatives to be developed and

sustained. These alternatives could include mangrove timber

production or development of new forms of aquaculture.

Additionally, restoration to the original habitat type may not

be the best social, economic, or ecological option. For

example, rather than restoring the area to a relatively

common ecosystem type, it may be preferable to restore not

to the original condition, but rather to a scarce type of habitat

within the ecosystem (Cairns 1988, Lewis Environmental

Services, Inc. and Coastal Environmental, Inc. 1996). This is

also known as “out-of-kind” restoration (Stevenson and

others 1999).

Figure 1—Diagram illustrating the continuum of options that are available for the rehabilitation of disused or

abandoned aquaculture ponds.

Restore the

environmental

conditions to reestablish a wetland

ecosystem.

Rehabilitate the

site to achieve an

alternative,

sustainable use.

Rehabilitate the

ponds so they can

be put back into

shrimp production.
24

Some areas may not be considered suitable for

restoration. In such cases, the best practical option may be

to re-establish shrimp farming in a well-managed and

sustainable manner. Similarly, the primary redevelopment

goal of coastal managers may be to recommence shrimp

farming in disused ponds.

In an area subject to storm activity, the primary goal may be

to restore natural coastal protection afforded by the buffer

function served by mangrove. Different goals are not

necessarily mutually exclusive. Rehabilitation of a degraded

mangrove can help to sustain shrimp aquaculture by acting

as a buffer to storm surges and reducing flooding of ponds.

It can help to provide nutrients to ponds as well as helping to

assimilate wastes drained from ponds.

The choice of options can be limited by the information

available on key factors that influence the costs and technical

difficulties of rehabilitating ponds. For example, ponds sited

in areas where Acid Sulfate Soils have been exposed to air

pose difficult problems of restoring neutral soil conditions

without costly liming or leaching the acids by flushing, which

can impose unacceptable impacts on neighboring rivers and

estuaries. There are fundamental gaps in the information

available on the environmental effects of leachate from Acid

Sulfate Soils and the persistence of chemicals, disinfectants,

and other materials used during shrimp culture. If new and

innovative treatments develop, then a variety of new

alternatives may be open to the coastal manager for

consideration (Stevenson and others 1999).

FACTORS THAT INFLUENCE OPTIONS FOR THE

REHABILITATION OF MANGROVE

Factors influencing the rehabilitation of mangrove were

reviewed by Chapman (1976), Field (1997), Hamilton and

Snedaker (1985), Lewis (1982, 1990a, 1990b), Lugo (1992),

Noakes (1951), Saenger and Siddiqi (1993), Siddiqi and

others (1993), and Watson (1928).

Some of the main factors that have a major influence on

both the sustainability of aquaculture and the rehabilitation

of disused sites include:

• the presence of Potential Acid Sulfate Soils (P.A.S.S.);

• erosion and subsidence of soils and associated

increased risks of flooding;

• waterlogging of soils; and

• the presence and longevity of antimicrobials and other

chemical compounds used in aquaculture operations.

There is a shortage of studies that quantify the

environmental conditions in disused shrimp ponds and their

effect on surrounding areas. Disused or abandoned ponds

or both may represent heavily degraded environmental

conditions that could prove technically difficult, expensive,

and time consuming to rehabilitate. In some case,

environmental changes brought about by developing ponds

in inappropriate situations may be irreversible. For example,

it is not possible to reverse the formation of Actual Acid

Sulfate Soils (A.A.S.S.) resulting from the oxidation of

Potential Acid Sulfate Soils (P.A.S.S.) caused by excavating

ponds. However, it may be possible to transform A.A.S.S. to

a pH neutral soil that has less acid-producing capacity.

Great care must be taken in choosing methods of dealing

with A.A.S.S. Methods commonly recommended by

agriculturists involve flushing the acids from the soil with

copious amounts of either fresh or brackish water, for

example, in Agency for Agricultural Research and

Development and the Land and Water Research Group

(1990). Although this may reduce the acidity in local areas, it

does not address the issue of the impact of acids draining

into adjacent streams, rivers, or estuaries. Field

observations by one of the authors in the Mekong Delta

indicated acid levels as low as pH 2.3 in canals dredged in

P.A.S.S. in an attempt to reclaim Melaleuca wetland forests

for agriculture. P.A.S.S. can extend over large areas and

represents a very large reservoir of potential acid materials.

Flushing of acids from pond sites may in the short term

appear to be a viable solution. However, unless water levels

are maintained above the P.A.S.S. layers, oxidation will

continue with the result of continuing leaching of acids.

Maintaining water levels in the soil profile can be very

difficult where groundwater is being extracted to irrigate

crops or to moderate salinity and temperatures in

aquaculture sites.

With the clearance of trees and other vegetation to create

ponds, surface erosion and subsidence and compaction of

the soil profile can take place. This can have a series of

knock-on effects, including increased risks of flooding from

river systems as well as increased vulnerability from oceanic

storm surges. These conditions are very difficult to reverse.

Additionally, it is difficult to counteract the adverse effects on

surrounding areas resulting from clearance of mangrove and

other wetland forest types. Lahmann and others (1987)

discuss the impact of shrimp aquaculture siting in basin

mangrove forests and “salitrales” (salt flats) in the southern

Gulf of Guayaquil, Ecuador. The basin forests are often

characterized as “unproductive” in major shrimp aquaculture

countries like Thailand, but Lahmann and others (1987)

point out that the local“...declining abundance of shrimp

postlarva in Ecuadorian estuaries...” may at least in part be

due to “...the disproportionate elimination of sources of

dissolved organic matter...” and “...may be the dominant

cause of the reduction in wild shrimp postlarva stocks...” (p.

242).

There is evidence that mangrove forests around the world

can recover from a degraded or heavily harvested condition

if: (a) normal freshwater flows entering the mangrove are not

disrupted, (b) the normal tidal hydrology is not disrupted,

and (c) the availability of waterborne seeds or seedlings

(propagules) from adjacent mangrove stands is maintained.

Through afforestation, mangroves can also be established

on unvegetated, intertidal flats where they would not

normally grow. These areas, however, are limited in extent

and often serve other ecological purposes such as feeding

areas for wading birds like herons and egrets. They may also

support valuable submerged aquatic vegetation like

seagrass meadows that are a valuable marine habitat in

their own right (Phillips and McRoy 1980).

Based on the ability of mangrove to restore itself, it has been

recommended that plans for restoration should first look at

the potential existence of stresses such as blocked tidal
25

inundation, which might prevent secondary succession

from occurring and to remove those stresses before

attempting additional restoration (Hamilton and Snedaker

1985, Lugo 1992). The identification of stresses and

evaluation of their effects on natural regeneration are not

simple tasks. Studies in Panama (Duke 1996) and

Indonesia (Soemodihardjo and others 1996) report

successful natural regeneration of mangrove subjected to

stresses from oil spills and logging, respectively. In

Panama, Duke (1996) observed that “...densities of natural

recruits far exceeded both expected and observed

densities of planted seedlings in both sheltered and

exposed sites...” (p. 228). Soemodihardjo and others

(1996) report that only 10 percent of a logged area in

Tembilahan, Indonesia, needed replanting because “...the

rest of the logged over area...had more than 2,500 natural

seedlings per hectare” (p. 109). In cases where natural

recovery is not occurring through natural seedling

recruitment, assisting natural recovery through planting is

one option.

Unfortunately, many mangrove restoration projects move

immediately into planting of mangroves without determining

if natural recovery is taking place and which stresses may

prevent natural regeneration and the likely effectiveness of

replanting. This often results in major failures of planting

efforts (Stevenson and others 1999). For example, Sanyal

(1998) has recently reported that between 1989 and 1995,

over 9000 ha of mangroves were planted in West Bengal,

India, with only a 1.52-percent success rate. The World Bank

funded Central Visayas Regional Project-I, Nearshore

Fisheries Component in the Philippines that targeted 1000

ha for mangrove planting between 1984–92. An evaluation of

the success of the planting in 1995–96 (Silliman University

1996) indicated only 18.4 percent of 2,927,400 mangroves

planted over 492 ha had survived. Another planned 30 000-

ha planting effort in the Philippines (Fisheries Sector

Program, 1990–95) funded by a $150 million loan from the

Asian Development Bank and the Overseas Economic

Cooperation Fund of Japan was cut short after only 4792 ha

were planted due to similar problems (Ablaza-Baluyut 1995).

Restoration of Disused Aquaculture Ponds

to Mangrove Forests

Detailed studies by Lewis and Marshall (1997) of pond

restoration at field sites in Central America and the

Philippines indicate that the most important factor in

successful restoration of ponds to mangroves is the reestablishment of the tidal hydrology to the maximum extent

possible. Restoration of mangroves through restoring the

natural hydrology has been emphasized before by Hamilton

and Snedaker (1984), Lugo (1992), and by Olsen and

Arriaga (1989). Turner and Lewis (1996) also give examples

of successful restoration of mangroves through restored

hydrology alone.

Lewis and Marshall (1997) have suggested five critical steps

to achieve successful mangrove restoration in general and

pond restoration in particular:

• understand the autecology (single species ecology) of

the mangrove species at the site, in particular the

patterns of reproduction, propagule distribution, and

successful seedling establishment;

• understand the normal hydrologic patterns that control

the distribution and successful establishment and growth

of targeted mangrove species;

• assess the modifications of the previous mangrove

environment that occurred and that currently prevent

natural secondary succession;

• design the restoration program to initially restore the

appropriate hydrology and utilize natural volunteer

mangrove propagule recruitment for plant establishment;

and

• only utilize actual planting of propagules, collected

seedlings, or cultivated seedlings after determining,

through steps 1 through 4, that natural recruitment will

not provide the quantity of successfully established

seedlings, or rate of stabilization, or rate of growth of

saplings, established as goals for the restoration project.

These critical steps are often ignored, and failure in most

restoration projects can be traced to proceeding directly to

step 5 without considering steps 1 through 4. Lewis and

Marshall (1997) refer to this approach as “gardening”, where

simply planting mangroves is seen as all that is needed.

SELECTING APPROPRIATE REHABILITATION

OPTIONS

The selection of the most appropriate options for

rehabilitation depends heavily upon local as well as national,

economic, social, and environmental priorities. From an

economic viewpoint, assessing the value of the flows of

resources derived from the former ecosystem provides an

important benchmark against which to assess the relative

costs and benefits that can be derived from different

rehabilitation options. Such benchmarks should incorporate

both the tangible economic goods and services as well as

the less tangible environmental goods and services

provided by the wetland forest or other ecosystems before

they were subjected to conversion. These environmental

goods and services include flood and storm protection,

sediment and toxicant removal, erosion mitigation, and

nutrient export.

Of particular importance are the benefits provided by

mangrove in coastal protection from typhoons or hurricanes.

In areas subject to strong storm activity for West Bengal,

Bangladesh, and Mozambique, the removal of mangrove to

form shrimp ponds has made coastal activities more

vulnerable to storm surges and flooding. In some states in

India, e.g., Andrha Pradesh, the buffering capacity and the

erosion mitigation provided by mangrove is particularly

important because the costs associated with the

construction of artificial structures to combat erosion and to

protect from storm damage can total more than.$12,000

(U.S.) per meter. In such areas, the case for restoration may

be very strong, and the costs involved in restoration may be

small when compared with the costs of constructing these

artificial structures. The benefits in such cases would be

both financial, in terms of costs avoided, and social, in terms

of the avoidance or minimization of the loss of human life

and property.

There are well-established economic techniques for

undertaking such analyses. For example, see section 6 of

the Mangrove Area Management Handbook (Hamilton and
26

Snedaker 1985). Mangroves are regarded as particularly

rich in terms of the goods and services they provide

(Burbridge 1990, Dixon 1989, Hamilton and Snedaker 1984)

and some researchers have made estimates of the

economic benefits of mangroves both in terms of fisheries

and other benefits. There is less information available on the

economic value of other land-use types or other habitats in

tropical regions, e.g., salt pans, tidal swamp forests,

Melaleuca wetlands, or mud flats.

However, well-constructed economic cases that include a

full range of social as well as environmental factors for

different rehabilitation options are seldom undertaken.

Justifications for rehabilitation such as those by Sidall and

others (1985) who state with reference to Ecuador, Panama,

and the Philippines that “reclamation of abandoned ponds

should be encouraged... and poorly sited or engineered

ponds should be breached to promote eventual

recolonization” would benefit from rigorous assessment of

the benefits of different rehabilitation options. This also

applies to justifications for rehabilitating general aquaculture

related degradation in mangroves. For example, Ishwaran

(1996) argues that given the importance of mangroves,

there is a need to rehabilitate degraded areas through

planting and the introduction of environmentally friendly

aquaculture technology.

The case for rehabilitating disused shrimp ponds should

also be viewed objectively. The direct loss of shrimp sales

revenue from ponds, which fail due to poor management,

can be in the region of.$15,000 to $25,000 (U.S.) per

hectare per year. The costs of the system degrading further

and posing a risk to neighboring habitats or land-use types

should also be taken into consideration. Damaged resources

are often unstable and actively deteriorating, and, in general,

if deterioration is not arrested, repair may become

progressively more expensive and difficult; i.e.,

redevelopment costs must be balanced against ‘costs

avoided’ (Stevenson and others 1999).

There can be added benefits from the rehabilitation of

degraded pond sites to an alternative income-generating

activity. For example, by redeveloping an area that has

become degraded and helping stakeholders to regain

productive use of the sites, pressure will be reduced on

neighboring areas, which are perceived to be of high value

for conversion and are at risk of degradation. The resulting

re-establishment of production of useful goods and services

in a rehabilitated or restored area may serve to help

maintain the flows of economic and environmental goods

and services provided by undisturbed ecosystems. This will

allow these systems to continue to sustain other forms of

human activity, as well as meeting international obligations

such as maintaining biological diversity.

Where ponds were converted from other land-use types,

such as fish ponds, paddy fields, or other agricultural lands,

socio-economic reasons may exist for restoring ponds to

their prior uses, particularly where local reliance on

subsistence agriculture or aquaculture exists. However,

there is little in the literature that supports restoration to prior

land-use types. Where restoration to a ‘natural’ habitat (such

as mangrove) is considered inappropriate, technically very

difficult, or too expensive, then restoration to prior or other

land-use types may be considered. This may represent the

best option in terms of economic feasibility, environmental

acceptability, and maximal sustainable productivity

(Stevenson and others 1999).

Where ponds have been constructed in inappropriate

locations, such as those frequently hit by typhoons or

hurricanes, there is a strong case for not rehabilitating ponds

to alternative uses that also will be vulnerable and difficult to

sustain. In cases where areas which are not normally hit by

hurricanes but experience El Niño years, for instance, along

the Pacific Coast of Latin America and the United States,

there can be a case for rehabilitation (or even restoration) of

ponds if the benefits outweigh the risks. In such cases,

management practices can reduce the vulnerability of

aquaculture production to periodic events.

Selection of Rehabilitation Options and

Their Implementation in Practice

There are limited references to the practical restoration of

abandoned shrimp ponds. Given the extensive areas of

mangrove converted unsuccessfully to shrimp ponds, there

have been surprisingly few reports of attempts to restore

aquaculture ponds back to mangroves. Anecdotal reports of

up to 13 000 ha of ponds restored in Thailand and several

thousand hectares in Vietnam have been noted by

Stevenson and others (1999). It is significant to note that

Field (1997), in compiling and editing reports on mangrove

restoration from 13 countries including Thailand, Malaysia,

Vietnam, Indonesia, India, Pakistan, and Bangladesh, does

not report a single occasion where pond restoration was

attempted. In fact, the only mention of aquaculture ponds in

this work is by Aksornkoae (1997), in which mangroves were

restored between existing shrimp aquaculture ponds in

Pattani Province, Thailand. In several areas in Thailand (and

probably elsewhere), mangrove has been planted in bund

walls and in areas adjacent to ponds as part of the

aquaculture management system in order to stabilize

sediments and to improve water quality.

Stevenson and others (1999) observed that pond

rehabilitation is an ongoing concern, and there are probably

many localized efforts to either reforest ponds in former

mangrove areas or to put disused ponds to alternative uses.

However, the data that do exist are often of poor quality, or

they are poorly disseminated and difficult to obtain or verify.

Consequently, it is not possible to draw any substantive

conclusions from them. The authors concluded that in most

cases, evaluation procedures or assessment of pond

condition do not take place prior to the initiation of pond

rehabilitation projects. Therefore, the reasons behind either

the success or failure of pond rehabilitation projects are not

known, and there is no ‘learning curve’ or lessons learned

from these endeavors. This is not a rational way to continue

(Stevenson and others 1999).

CONCLUSIONS

Aquaculture forms an important activity in many of the

world’s coastal regions, and its importance as a source of

income, employment, and exports is likely to continue to

expand for the foreseeable future. Aquaculture of a generally
27

nonintensive nature has been a part of coastal land and

water use for many centuries in Asia, and it has proven

sustainable. However, the rapid expansion of poorly planned

and managed semi-intensive and intensive shrimp

aquaculture has created a number of significant adverse

environmental, economic, and social effects. In turn, shrimp

aquaculture has often been adversely affected by impacts

from other forms of human development.

The combined effects of poor standards of aquaculture

development and adverse impacts on aquaculture

operations have led to unnecessary destruction of coastal

wetland forests, nonsustainable aquaculture production,

disuse of pond sites, and even abandonment of the land and

the loss of the investment of peoples’ labor and capital.

Reluctance on the part of the aquaculture industry and

government to admit that shrimp aquaculture has often

proven nonsustainable has helped to disguise the extent of

unproductive shrimp farming. However, extensive areas are

believed to be involved in many of the poorer developing

nations such as India, the Philippines, and Indonesia.

Developing nations cannot afford to allow the extensive

areas of nonsustainable, unproductive shrimp farms to

remain idle if they are to meet the development needs of

coastal communities, as well as national development

objectives and international obligations such as the

conservation of biodiversity. Substantial scientific effort is

required to analyze the factors that lead to nonsustainable

aquaculture and to help the 50 or more shrimp-producing

nations find ways to rehabilitate unproductive, disused, and

abandoned areas. This will serve both to help people to

develop sustainable uses of coastal areas and resources

already committed to aquaculture and to reduce pressures

on remaining coastal ecosystems and the renewable

resources they generate.

Options for the rehabilitation of areas that can no longer

sustain shrimp production need to be identified, tested,

evaluated, and demonstration projects established as a

means of engaging the interest and active support of

stakeholders. Efforts to improve the dialogue between

donors, researchers, aquaculturists, and governmental

bodies will be essential in developing workable solutions.

Unfortunately, much of the research into pond rehabilitation

carried out to date has been conducted without adequate

site assessment, without documentation of the

methodologies or approaches used, and often lacks

subsequent follow up or evaluation. Those projects, which

have not been successful, are rarely documented and

information on them is largely anecdotal and hard to obtain.

The reasons for success or failure are still largely

guesswork, and we are still at the beginning of what may

prove to be a steep ‘learning curve.’

To reduce the length of time required to address this

learning curve and avoid further unnecessary damage to

wetland forests, consideration should be given by the

international scientific community and donors to placing

greater emphasis upon working with developing nations

where these issues are acute and where poverty and lack of

effective development assistance will drive people to

degrade more areas through repeating mistakes of the past.

The potentially adverse effects of new aquaculture

development can generally be avoided through good

planning and management. To be fully successful, such

plans and management arrangements must recognize that

aquaculture should have equal rights of access to and use of

natural resources and a good quality environment. It is

suggested that Integrated Coastal Zone Management can

provide a beneficial framework for the development of

aquaculture where due care and attention are given to the

maintenance of the functional integrity of coastal

ecosystems that sustain aquaculture and other natural

resource-dependent activities. It is also suggested that

multiple use management of coastal ecosystems will provide

a better basis for integrating aquaculture with other activities,

which have a common dependence on the functions and

resources provided by one or more coastal ecosystems.

ACKNOWLEDGMENTS

This paper draws heavily on papers prepared by the first

author in collaboration with N.J. Stevenson, R.R. Lewis, A.J.

Edwards, and S. Jennings, and we wish to acknowledge the

value of their work in supporting this paper.

REFERENCES

Anon. 1985. The state of food and agriculture 1984. FAO Agric. Ser.

18. Rome: Food and Agriculture Organization of the United

Nations. 185 p.

Ablaza-Baluyut, E. 1995. The Philippines fisheries sector program.

In: Proceedings of a workshop on coastal and marine environmental management; 1995 March 27–29; Bangkok, Thailand.

Bangkok, Thailand: Manila, Philippines: Asian Development

Bank: 156–177.

Agency for Agricultural Research and Development and the Land

and Water Research Group. 1990. Workshop on acid sulphate

soils in the humid tropics; 1990 November 20–22; Bogor,

Indonesia. 392 p.

Aksornkoae, S. 1997. Reforestation in mangrove forests in Thailanda case study in Pattani Province. In: Field, C.D., ed. Restoration

of mangrove ecosystems. Okinawa, Japan: International Society

for Mangrove Ecosystems: 53–63.

Barg, U.C. 1992. Guidelines for the promotion of environmental

management of coastal aquaculture development. FAO Tech.

Pap. 328. Rome: Food and Agriculture Organization of the United

Nations. 122 p.

Briggs, M.R.P.; Funge-Smith, S. 1994. Unsustainable shrimp culturecauses and potential solutions from experience in Thailand. In:

Development of strategies for sustainable shrimp farming. Rep.

to the Overseas Development Administration, Res. Pro. R4751.

App. 1. Stirling, UK: Institute for Aquaculture, Stirling University.

31 p.

Burbridge, P.R. 1990. Tidal wetland resources in the tropics. Journal

of Resource Management and Optimization. 7(1–4): 115–138.

Cairns, J. 1988. Restoration ecology: the new frontier. In: Cairns, J.,

ed. Restoration of damaged ecosystems. Boca Raton, FL: CRC

Press: 1–12. Vol. 1.
28

Chapman, V.J. 1976. Mangrove vegetation. Vaduz, Germany: J.

Cramer. 447 p.

Chou, L.M.; Chua, T.E.; Khoo, H.W. [and others], eds. 1991. Towards

an integrated management of tropical coastal resources. In:

International center for living aquatic resources management

(ICLARM): Conference proceedings 22; 1988 October 28–31;

Singapore. Manila, Philippines: International Center For Living

Aquatic Resources Management. 455 p.

Clements, F.E. 1928. Plant succession and indicators. New York:

H.W. Wilson Co. 453 p.

Dixon, J.A. 1989. Valuation of mangroves. Tropical Coastal Area

Management. 4(3): 2–6.

Duke, N. 1996. Mangrove reforestation in Panama. In: Field, C.D.,

ed. Restoration of mangrove ecosystems. Okinawa, Japan:

International Society for Mangrove Ecosystems: 209–232.

Fegan, D.F. 1996. Sustainable shrimp farming in Asia: vision or

pipedream. Aquaculture Asia. 1(2): 22–28.

Field, C.D., ed. 1997. Restoration of mangrove ecosystems.

Okinawa, Japan: International Society for Mangrove Ecosystems. 250 p.

Flaherty, M.; Karnjanakesorn, C. 1995. Marine shrimp aquaculture

and natural resource degradation in Thailand. Environmental

Management. 19(1): 27–37.

Gommes, R.; de Guerny, J.; Nachtergaele, F.; Brinkman, R. 1998.

Potential impacts of sea-level rise on populations and agriculture. Rome: Food and Agricultural Organization of the United

Nations.

Hambrey, J. 1996. Comparative economics in land use options in

mangrove. Aquaculture Asia. 1(2): 10–14.

Hamilton, L.S.; Snedaker, S.C., eds. 1984. Restoration and

establishment. Sec. 111. In: Handbook of mangrove area

management. Honolulu: East West Center, International Union

for the Conservation of Nature, United Nations Education Social

and Cultural Organization (UNESCO), United Nations Development Program (UNDP): 102–108.

Hodgeson, G.; Dixon, J.A. 1988. Logging versus fisheries and

tourism in Palawan. Pap. 7. Honolulu: East West Environment

and Policy Institute. 95 p.

Holdgate, M.W. 1993. The sustainable use of tropical coastal

resources-a key conservation issue. Ambio. 22: 481–482.

ICES. 1995. Principles and practical measures for the interaction of

mariculture and fisheries in coastal area planning and management; Rep.: Working Group on Environmental Interactions of

Mariculture. Copenhagen, Denmark: Mariculture Committee;

International Council for Exploration of the Sea. 28 p.

ICES. 1996. Report: working group on environmental interactions of

mariculture. Copenhagen, Denmark: Mariculture Committee;

ICES CM 1996; International Council for Exploration of the Sea.

ICES. 1997. Report: working group on environmental interactions of

mariculture. Copenhagen, Denmark: Mariculture Committee;

ICES CM 1997; International Council for Exploration of the Sea.

ICES. 1998. Report: working group on environmental interactions of

mariculture. Copenhagen, Denmark: Mariculture Committee;

ICES CM 1998; International Council for Exploration of the Sea.

ICES. 1999. Report: working group on environmental interactions of

mariculture. Copenhagen, Denmark: Mariculture Committee;

ICES CM 1999; International Council for Exploration of the Sea.

Ishwaran, N. 1996. Global concepts and local practice: UNESCO

experience in implementing the biosphere reserve concept and

its relevance to integrated coastal resource management. In:

Phan Nguyen Hong, Nataragjan Ishwaran, eds. Proceedings of

the ECOTONE V seminar on community participation in

conservation, sustainable use and rehabilitation of mangrove in

South East Asia; 1996 January 8–12; Ho Chi Minh City, Vietnam.

Ho Chi Minh City, Vietnam: Agricultural Publishing House of

Vietnam: 221–228.

Jayasinghe, J.M.P.K. 1995. Sri Lanka. In: FAO/NACA report on

regional study and workshop on the environmental assessment

and management of aquaculture development (TCP/RAS/2253).

Network of Aquaculture Centers in Asia and the Pacific (NACA),

Environment and Aquaculture Development Series 1. Bangkok,

Thailand: Network of Aquaculture Centers in Asia and the

Pacific; Food and Agriculture Organization of the United Nations:

357–376.

Juliano, R.; Anderson, J.; Librero, A.R. 1981. Coastal zone management in the Philippines. In: Collier, W.; Chia, L.S.; Soysa, C., eds.

Man, land, and sea. Bangkok, Thailand: Agricultural Development Council: 219–240.

Kunstadter, P.; Bird, E.C.F.; Sabhasri, S., eds. 1985. Proceedings of

the workshop on man in the mangroves. Tokyo, Japan: United

Nations University. 117 p.

Lahmann, E.J.; Snedaker, S.C.; Brown, M.S. 1987. Structural

comparisons of mangrove forests near shrimp ponds in southern

Ecuador. Interciencia. 12(5): 240–243.

Landesmann, L. 1994. Negative impacts of coastal aquaculture

development. World Aquaculture. 25(2): 12–17.

Lewis Environmental Services, Inc. and Coastal Environmental, Inc.

1996. Setting priorities for Tampa Bay habitat protection and

restoration: restoring the balance. St. Petersburg, FL: Tampa Bay

National Estuary Program. 106 p.

Lewis, R.R. 1982. Mangrove forests. In: Lewis, R.R., ed. Creation

and restoration of coastal plant communities. Boca Raton, FL:

CRC Press: 153–172.

Lewis, R.R. 1990a. Creation and restoration of Coastal Plain

wetlands in Florida. In: Kusler, J.A.; Kentula, M.E., eds. Wetland

creation and restoration: the status of the science. Washington,

DC: Island Press: 73–101.

Lewis, R.R. 1990b. Wetlands restoration/creation/enhancement

terminology: suggestions for standardization. In: Kusler, J.A.;

Kentula, M.E., eds. Wetland creation and restoration: the status

of the science. Washington, DC: Island Press: 417–422.

Lewis, R.R. 1992. Coastal habitat restoration as a fishery management tool. In: Stroud, R.H., ed. Stemming the tide of coastal fish

habitat loss: Proceedings of a symposium on conservation of

coastal fish habitat; 1991 March 7–9; Baltimore. Savannah, GA:

National Coalition for Marine Conservation, Inc.: 169–173.

Lewis, R.R.; Kusler, J.A.; Erwin, K.L. 1995. Lessons learned from

five decades of wetland restoration and creation in North

America. In: Montes, C.; Oliver, G.; Molina, F.; Cobos, J., eds.

Bases ecologicas para la restauracion de humedales en la

Cuenca Mediterranea: Proceedings of a meeting; 1993 June 7–

11; University of La Rabida, Spain. Junta de Andaluca, Spain:

Andaluca: 107–122.
29

Lewis, R.R.; Marshall, M.J. 1997. Principles of successful restoration of shrimp aquaculture ponds back to mangrove forests

[Abstract]. Programa/resumes de Marcuba ’97; 1997 September

15–20; La Habana, Cuba. Palacio de Convenciones: 126.

Lugo, A.E. 1992. Restoration of mangroves. In: Thayer, G.W., ed.

Restoring the nation’s marine environment. College Park, MD:

Maryland Sea Grant Program. 716 p.

Lundin, C.G.; Linden, O. 1993. Coastal ecosystems: attempts to

manage a threatened resource. Ambio. 22: 468–473.

Macintosh, D.J. 1996. Mangroves and coastal aquaculture; doing

something positive for the environment. Aquaculture Asia. 1(2):

3–8.

Noakes, D.S.P. 1951. Notes on the silviculture of the mangrove

forest of Matang, Perak. Malay Forester. 14: 183–196.

Office of Environmental Policy and Planning (OEPP), Ministry of

Science and Technology, Thailand. 1994. The environmental

management of coastal aquaculture: a study of shrimp culture in

Samut Sakhon and Chantaburi Provinces. Bangkok, Thailand:

Network of Aquaculture Centers in Asia-Pacific. 195 p.

Ogburn, D.M.; Ogburn, N.J. 1994. Intensive pond culture trials of the

green grouper (Epinephelus malabaricus Bloch et Schneider) in

the Philippines. In: Chou, L.M.; Munro, A.D.; Lam, T.J. [and

others]. The third Asian fisheries forum 1; 1992 October 26–30;

Singapore. Manila, Philippines: Asian Fisheries Society: 74–77.

Olsen, S.; Arriaga, L. 1989. Establishing a sustainable shrimp

mariculture industry in Ecuador. Narragansett, RI: The University

of Rhode Island, Coastal Resources Center; Washington, DC:

U.S. Agency for International Development. 139 p.

Phillips, R.C.; McRoy, C.P, eds. 1980. Handbook of seagrass biology:

ecosystem perspective. New York: Garland STPM Press. 353 p.

Potaros, M. 1995. Country report on Thailand. Annex II–16. In: FAO/

NACA 1995 regional study and workshop on the environmental

assessment and management of aquaculture development

(TCP/RAS/2253). Network of Aquaculture Centers in Asia and

the Pacific (NACA), Environment and Aquaculture Development

Series 1. Bangkok, Thailand: Network of Aquaculture Centers in

Asia and the Pacific: 377–390.

Primavera, J.H. 1995. Mangroves and brackishwater pond culture in

the Philippines. Hydrobiologia. 295: 303–309.

Primavera, J.H.; Agbayani, R.F. 1996. Comparative strategies in

community based mangrove rehabilitation programs in the

Philippines. In: Proceedings of the ECOTONE V regional

seminar: community participation in conservation, sustainable

use and rehabilitation of mangroves in Southeast Asia; 1996

January 8–12; Ho Chi Minh City, Vietnam. Ho Chi Minh City,

Vietnam: Mangrove Ecosystem Research Centre (MERC),

Vietnam National University: 34.

Saenger, P.; Siddiqi, N.A. 1993. Land from the sea: the mangrove

afforestation programme of Bangladesh. Ocean and Coastal

Management. 20: 23–29.

Sammut, J.; Mohan, C.V. 1996. Processes and impacts of estuarine

acidification: the Richmond River, NSW and Karnataka, India. In:

One-day symposium on issues for eastern Australia and Asia:

Acid sulphate soils-causes, management, effects, rehabilitation,

prevention; 1996 July 29; Sydney, Australia. Sydney, Australia:

University of Sydney, Earth Sciences Foundation, Department of

Geology and Geophysics. 4 p.

Sanyal, P. 1998. Rehabilitation of degraded mangrove forests of the

Sunderbans of India [Abstract]. Program of the international

workshop on the rehabilitation of degraded coastal systems;

1998 January 19–24; Phuket, Thailand. Phuket, Thailand: Phuket

Marine Biological Center: 25.

Sidall, S.E.; Atchue, J.A.; Murray, R.L., Jr. 1985. Mariculture

development in mangroves: a case study of the Philippines,

Ecuador, and Panama. In: Clark, J.R., ed. Coastal resources

management: development case studies. Renewable Resources

Information Series, Coastal Management Publ. 3. Columbia, SC:

Research Planning Institute: 1–53. [Prepared for the U.S.

Department of the Interior, National Park Service, and the U.S.

Agency for International Development].

Siddiqi, N.A.; Islam, M.R.; Khan, M.A.S.; Shahidullah, M. 1993.

Mangrove nurseries in Bangladesh. Occas. Pap. 1. Okinawa,

Japan: International Society for Mangrove Ecosystems. 15 p.

Silliman University. 1996. Assessment of the Central Visayas

Regional Project-I: nearshore fisheries component. Dumaguete

City, Philippines: Silliman University. 126 p. Vol. 1.

Soemodihardojo, S., Wiroatmodjo, P.; Mulia, F.; Harahap, M.K. 1996.

Mangroves in Indonesia-a case study of Tembilahan, Sumatra.

In: Fields, C.D., ed. Restoration of mangrove ecosystems.

Okinawa, Japan: International Society for Mangrove Ecosystems: 96–110.

Sreng, L. 1996. Mangroves in Kampot Province, Cambodia. In:

Proceedings of the ECOTONE V regional seminar: community

participation in conservation, sustainable use and rehabilitation

of mangroves in Southeast Asia; 1996 January 8–12; Ho Chi

Minh City, Vietnam. Ho Chi Minh City, Vietnam: Mangrove

Ecosystem Research Centre (MERC), Vietnam National

University: 13.

Stevenson, N.J. 1997. Disused shrimp ponds: options for redevelopment of mangrove. Coastal Management. 25(4): 423–425.

Stevenson, N.J.; Lewis, R.R.; Burbridge, P.R. 1999. Disused shrimp

ponds and mangrove rehabilitation. In: Streever, W.J., ed. An

international perspective on wetland rehabilitation. Dordrecht,

The Netherlands: Kluwer Academic Publishers: 277–297.

Tuan, M.S. 1996. Building up the strategy for mangrove management in Vietnam. In: Proceedings of the ECOTONE V regional

seminar: community participation in conservation, sustainable

use and rehabilitation of mangroves in Southeast Asia; 1996

January 8–12; Ho Chi Minh City, Vietnam. Ho Chi Minh City,

Vietnam: Mangrove Ecosystem Research Centre (MERC),

Vietnam National University. 35 p.

Turner, R.E.; Lewis, R.R. 1997. Hydrologic restoration of coastal

wetlands. Wetlands Ecology and Management. 4(2)(Special):

65–72.

Watson, J.G. 1928. Mangrove forests of the Malay Peninsula. Rec.

6. Malayan Forester. 276 p.

Yap, W.G. 1997. Can the Philippines produce enough fish for the

multitude? Aquaculture Asia. 2(3): 32–38.
30

EFFECTS OF HYDROLOGIC CONDITIONS ON

BIOGEOCHEMICAL PROCESSES AND ORGANIC POLLUTANT

DEGRADATION IN SALT MARSH SEDIMENTS

W. James Catallo1

Abstract—This work addressed the influence of tidal vs. static hydrologic conditions on biogeochemical processes and the

transformation of pollutant organic chemicals (eight representative N-, O-, and S-heterocycles (NOSHs) from coal

chemicals, crude oils, and pyrogenic mixtures) in salt marsh sediments. The goals were to: (1) determine the effects of static

(flooded, drained) vs. dynamic (tidal) hydrology on redox potential (Eh) dynamics, trace gas evolution, and pollutant

transformation; (2) deploy hydrodynamic microcosms for this purpose that were reproducible, well controlled, and

adequately monitored; and, (3) develop analytical approaches for target pollutant chemicals that allowed for detection of

small but significant concentration differences between time points and treatments, i.e., isotopic dilution. NOSH-amended

sediments were exposed to three hydrologic conditions: static drained (oxidized redox potentials), static flooded (reduced

redox potentials), and diurnal-tidal (alternating redox potential). The rate of NOSH transformation and the number of NOSHs

degraded decreased in the following order: drained = tidal flooded. This indicated that sediments and associated biota

exposed to tidal pulsing removed more NOSH compounds faster and to lower levels than flooded, highly reducing

sediments.

1

 Associate Professor of Environmental Toxicology, Laboratory for Ecological Chemistry, School of Veterinary Medicine, Louisiana State

University, Baton Rouge, LA 70803.

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability of

wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.

INTRODUCTION

The work summarized here examined the influence of

hydrology on biogeochemical processes that govern

transport and transformation of selected pollutant chemicals

in wetland sediment-plant systems. “Biogeochemical” refers

to a group of coupled biological, e.g., microbial, plant and

physicochemical, e.g., flood-drain events from tides and

floods, processes that to a large extent determine the

productivity, habitat quality, and regional-global significance

of wetland and other ecosystems (Catallo 1999, Catallo and

others 1999). Previous microcosm studies have confirmed

that degradation and transformation rates and pathways of

aromatic hydrocarbons (AHs) and N-, O-, and Sheterocycles (NOSHs) differed significantly in marine

sediments of different particle sizes and under oxidized vs.

reduced conditions (Catallo 1996b, Catallo and Gambrell

1994). In the cited work, AH and NOSH transformation was

evaluated in stirred, bubbled, controlled Eh/pH reactors

containing sediment slurries maintained as “oxidized,

aerobic,” “moderately reducing, anoxic,” and “highly reducing,

methanogenic.” AH and NOSH degradation rates generally

were: oxidized moderately reducing >> methanogenic. The

reactors used were first-order feedback control systems

whose output domains were predefined by desired Eh

ranges for each treatment. As a result, Eh vs. time outputs

were sinusoidal. Fourier spectral analysis of the Eh

waveforms demonstrated that periodic sampling

interventions and other system perturbations were detected

by the electrodes and preserved in the time series for each

system. This suggested that the Eh might change rapidly

and reproducibly in response to other periodic forcing, such

as diurnal tidal flushing, but this had not been well

documented. Further, it was clear that the terms “oxidized,

aerobic” applied to wetland sediments were something of an

oxymoron. Nevertheless, this treatment type was evaluated

because large quantities of wetland and aquatic sediments

are disposed of in upland or otherwise well-drained

situations, so evaluation of these sediments was deemed

important, albeit not especially compelling from the

perspective of “wetlands ecology.”

Apart from the differential NOSH and AH transformation

rates observed in these experiments, results strongly

suggested that measured redox potential (Eh) can be a

dynamic nonlinear variable in wetland sediments that can

respond quickly and significantly to changing conditions, and

that this information can be preserved in a time series of Eh

provided adequate sampling intervals are employed. As Eh

is considered a biogeochemical “master variable” in many

ecosystems, the behavior of this variable should be

adequately and accurately represented by the data at all

scales relevant to the experimental design (Catallo 1993,

Catallo and others 1999, Lindsay 1991). It was clear that the

interaction of tidal water level changes with sediments

should be included as a treatment in studies of

biogeochemical processing of contaminants. In order to do

this, effort was expended in the design, monitoring, and

evaluation of hydrodynamic sediment microcosms (20 gal)

and, later, mesocosms (300 to 3000 gal) containing a

biocenosis comprised of sediments, plants, microbes,

meiofauna, and macrofauna (Catallo 1999).

The central aims of the current study were: (1) determine the

effects of static vs. dynamic hydrology on biogeochemical

≤
31

behavior of salt marsh sediments, particularly with respect to

major output variables, e.g., Eh dynamics, trace gas

signatures, and pollutant transformation; (2) design, build,

and optimize microcosms for this purpose that are

reproducible, well controlled, and adequately monitored in

the laboratory or greenhouse; and, (3) develop extraction/

analytical approaches for NOSH target chemicals that allow

for detection of small but significant concentration

differences between time points and treatments. This last

item was important because most common methods used

for identification of chemicals in sediments are

semiquantitative, i.e., provide for order-of-magnitude

differentiation in analyte concentrations, and it was expected

that differences between NOSH transformation profiles in

different treatments of this study would be below this

threshold (Catallo 1996a).

The pollutant chemicals selected for this work included

representative N-, O-, and S-heterocycles (NOSHs) that

occur in coal chemicals, several crude oils, and pyrogenic

(combustion-generated) mixtures (fig. 1) (Turov and others

1987). These and other NOSHs are found in large quantities

in pollutant mixtures near certain industrial activity,

hazardous waste sites, and major harbors. They enter

coastal marine and wetland ecosystems through: (a) direct

dumping/spillage; (b) fluvial transport of dissolved and

sediment-associated chemicals from source outfalls,

deposition of combustion-generated airborne particles,

semivolatiles, e.g., from the so-called in situ burns of spilled

oils, and polar residues; and, (c) assorted natural processes,

including marine oil seeps and reactions between biogenic

organic matter and sub- and supercritical water in various

geochemical settings. They are of continuing interest

because they can accumulate to bioavailable levels in

aquatic sediments, and many NOSH compounds are acutely

toxic, with some, e.g., the quinolines, carbazoles, and

benzoquinolines and their oxidized metabolic and

environmental transformation products, being mutagens or

carcinogens or both (Catallo 1996b, Catallo and others

1994, Warshawsky 1992).

MATERIALS AND METHODS

Hydrodynamic (Tidal) Wetland Microcosms

The microcosms for this work were comprised of 20 gal

glass aquaria containing sediment columns as shown in

figure 2. Sediment samples were collected from a

streamside S. alterniflora salt marsh site in Terrebonne

Parish, LA. The sediments were mixed with quartz sand and

commercial seed starter (5 percent w/w) and then loaded

into cylinders (26 cm i.d.) made first of wire, and ultimately of

plastic mesh (0.5 cm) to a height of 38 cm. The wire mesh

was discontinued after preliminary runs because it interfered

with Eh measurements. The systems were covered with

plexiglass containing throughputs for water lines, gas

sampling, and electrode wire leads, and sealed with silicone

plastic. A diurnal tide was simulated by pumping artificial

seawater (15 g/L Instant Ocean) into and out of the

aquarium once daily, i.e., one high and one low per 24 hour,

using peristaltic pumps and controlled by battery-powered

timers. The seawater reservoir was continuously aerated. At

“high tide,” the sediment columns were covered with ca. 8 cm

of water. At “low tide” about 5 cm of water remained in the

aquarium. Two Eh electrodes (below) were positioned in the

top 1 cm of sediment (surface), and two additional

electrodes were placed deep within the column in the

continually flooded zone near the bottom. Trace gases were

analyzed by direct sampling of headspace gases under

vacuum to a 10-m cell connected to a Buck Scientific FTIR

spectrometer with gold optics and 4 per cm resolution.

Molecular sieve 5A and Ascarite A (Aldrich) traps were used

to reduce water and CO2

 spectral interference and increase

sensitivity. Compounds were identified using authentic gas

standards (Scott Specialty Gases) and published spectral

libraries from NIST. Digital background subtractions were

performed using spectra of normal laboratory air sampled

immediately prior to sampling the microcosms, and handled

identically; i.e., reference air samples were passed through

the sampling loop of the system and through the traps. Thus,

CO2 and water vapor signals in the experimental spectra

represent excesses of these materials in the microcosm

headspaces vs. the laboratory air.

 Quinoxaline 2-Methylquinoxaline 2,3-Dimethylquinoxaline

Phenothiazine Phenazine Dibenzothiophene

 Acridine Phenanthridine

Figure 1—NOSH target analytes examined in this study.

Figure 2—Schematic diagram of the tide simulation microcosm: (1)

data logger, (2) calomel reference electrode, (3) salt bridge, (4)

platinum electrodes, (5) concentrated KC1 solution, (6) soil column

enclosures, (7) water inlet/outlet, (8) automatic peristaltic pumps,

and (9) reservoir. Gas tight cover and FTIR feeds are not shown.
32

Redox potential (Eh) was logged using Pt wire–SCE

electrode cells and a multichannel logger (Catallo 1999).

The platinum electrodes were constructed by welding Cu

wire to 0.5 cm-lengths of 12 gauge Pt wire. The junction and

Cu lead were enclosed in glass tubes of varying lengths,

and both ends were sealed using a blowtorch. Hence, a ca.

0.5-cm length of Pt was exposed to the sediments whereas

junctions and leads were enclosed in sealed glass. The

electrodes were polished with light emory paper, soaked in

aqua regia and then distilled water. Each electrode was

calibrated prior to use with a 1-percent quinhydrone solution.

Samples were collected at rates between 1 Hz and 1 per

hour, depending on the application (in general, the rates

were 1 to 2 per hour).

Signal Acquisition and Analysis

A biogeochemical system can be viewed as a signal

processor, with inputs, e.g., light and heat cycles, water

fluctuations, inputs of organic matter and toxic chemicals,

impinging on the system, influencing its behavior, e.g.,

aerobic vs. anaerobic microbial metabolism, and influencing

the magnitude and time structure of numerous outputs

(water chemistry, trace gas identities, and evolution time

series). Shown schematically below is a generic system with

single continuous input and output variable functions, u(t)

and y(t),

respectively. The inputs mentioned are impulses or excitation

functions having importance to system processes, and all of

these must be known (identified) and measured accurately

with appropriate sampling frequencies. The former is

important because confounding and collinear variables can

invalidate normal statistical approaches and corrupt causal

ascriptions (Asher 1983, Catallo and others 1999) and the

latter because processes changing faster than the sampling

rate applied to them can cause error (aliasing) when spectral

analysis techniques are applied to the time series (Brockwell

and Davis 1991, Mallat 1998). These functions, (u(t)), can be

of many kinds, e.g., linear, spike, ramp, threshold, sinusoidal.

Signals are outputs, y(t) carrying information about system

processes. Signals are structured in time, whereas noise is

not, and increase relative to noise as the square root of the

sampling repetitions; (S/N)n

 = n1/2(S/N)1; where S is signal

amplitude, N is noise amplitude, (S/N)1 is signal:noise ratio

after one data acquisition (scan), n is the number of sample

scans, and (S/N)n

 is the signal:noise ratio after n scans. The

states of the system can be defined by the stable and

transient domains occupied by output signals y(t).

It frequently is assumed that biogeochemical processes

change slowly, i.e., weeks to months, and therefore

adequate sampling of many outputs can be performed at

rates on the order of 1 per day or longer (Catallo 1999,

Picek and others 2000). Unfortunately, it must first be

verified that there are no cyclic outputs from the system with

frequencies greater than the sampling rate, and this is rarely

done. As a result, the assumption that widely spaced data

points from dynamic systems can be connected by linear

interpolation to accurately reconstruct system temporal

behavior becomes a matter of faith rather than

demonstration. Further, attempts to analyze such a time

series in order to determine its component waves can be

highly inaccurate and misleading. The upshot of this is that

complex system variables (u(t)) and signals (y(t)) cannot be

interpreted accurately if data points are gathered at rates

slower than the changes in the process. For virtually all

ecological system processes, classical single resolution

Fourier analysis of time series suffers from major analytical

drawbacks, not the least of which being its inability to

accommodate nonstationarity in the data; i.e., transients and

other deviations from perfect statistical regularity in the

signals. Single resolution analysis of nonstationary time

series leads to aliasing of power spectra (Brockwell and

Davis 1991, Mallat 1998) in ways that are not readily

interpreted or identified. It was clear from Eh, pH,

temperature, and other time series from the microcosms that

nonstationarity was a major feature of the input and output

variables of concern in this work. As a result, multiresolution

“wavelet” analysis was applied to the time series of signals

from the microcosms.

Wavelet analysis is a recent development in applied

mathematics (Mallat 1998) that has stimulated an enormous

interest in many disciplines including signal processing/

reconstruction, vision science, medical imaging, and

physiology. The wavelet transform is a technique that allows

for the acquisition of both localized time and frequency

information about a signal. Its ability to dilate and shift time

windows with respect to a signal allows for transient features

to be accommodated, long- and short-term variations to be

compensated, and periodic features to be isolated,

identified, and, if needed, enhanced. In a sense, the wavelet

approach has operational and conceptual affinities with the

use of an oscilloscope to visualize waveforms in circuits:

time sampling windows are varied by the analyst in an

attempt to isolate component waveforms of complex signals

as standing waves, with frequency data available as the

inverse of the time window used to capture the particular

wave of interest. Waveform components and their alterations

by specific processors (distortion, clipping, chirping, and

damping) are obtained by thoroughly permuting the available

sampling windows.

The definition of the continuous time wavelet transform

(CWT) entails an “analysis wavelet,” ψ (t), and the family of

shifted and dilated versions

 ψ a,b = a-1/2

ψ (t –b)/a (1)

The transform of x (t) ∈ L2

(R) is then

 Tw[x] (a, b) = a-1/2 ∫ x(t) P *(t – b)/a dt (2)

The analysis wavelet, Ψ must satisfy the admissibility

condition

u (t) SYSTEM y (t)

SYSTEM

u(t) y(t)
33

 C p = ∫ Ψ(w)2

 /w dw < ∞ (3)

This condition guarantees the existence of the inverse

transform

x = Cψ-1 ∫ ∫ Tw [x] (a,b) ψ *(t – b)/a dadb;

 (intergrals evaluated ± ∞) (4)

One way to realize the transform is using the filter bank

interpretation and defining

 ha

 (t) = ψ (-t/a) (5)

It follows that

 Tw [x] (a,b) = x (t) ha (t) =∫ x(t) ha

 (b – t)dt (6)

So, for each scale, a, the CWT is comprised by the output of

a filter with impulse-response ha (t). As the filters are dilated

according to a scale parameter, their frequency response is

modified accordingly, at the discretion of the analyst. Hence,

the wavelet transform maps a function of one variable into a

function of two independent variables (time, scale). There is,

as a consequence, significant redundancy, and, if desired,

the time/scale parameters can be rendered discrete, the

signal preserved in the transform samples. A standard

discretization approach employs a logarithmic step for the

scale parameter and a scale-dependent discretization for the

temporal parameter. Thus a discretized wavelet transform

can be defined by

 Tw[m,n] = Tw[x] (2m, 2mn) (7)

In effect, a family of wavelets of the form ψm,n = 2-m/2ψ

(t / 2m –n) is deployed.

Even in the discretized case, the representation can be

optimized by requiring the wavelet family, ψm,n , to be an

orthonormal as a set. In this case for each value of the

integer m, the collection of vectors {ψm,n; n ∈Z}, defines a

subspace, Wm. As the dilation parameter is held constant, all

wavelets in that subspace have the same scale. The

projection of a function, x(t), on this subspace is interpreted

as the details of the function at that particular scale. Since

increasing the value of the integer, m, introduces larger

dilations in the basic wavelet, its time resolution decreases.

The sum of all details for all values m ≥m0 is said to give a

representation of x(t) up to resolution m0. Mathematically, the

representations exist in the subspace

 Vm = ⊕k≥m Wk (8)

One of the bulwarks of multiresolution analysis is that a

unique analysis function φ (t) can be defined such that the

collection {φ m,n = 2-m/2ψ(t / 2m –n)} spans the subspace, Vm,

and these are easy to compute recursively. If, for example,

the representation of a signal up to a fine resolution level,

m0

, is in the form

 Xmo = ∑ cmo

n φ mo,n 

(9)

one can define a digital filter, H, receiving as input the

discrete signal, { cmo= cmo

 n ; n ∈Z}, and producing as output

the coefficients of the next lower resolution, cmo+1

. Similarly,

the digital filter, G, can be defined as receiving the same

input and producing as output the details at resolution level

m0

.

The wavelet packet approach also applies to discrete time

functions such as those obtained by sampling a continuous

time signal. It can be considered as the discrete version of

the multiresolution decomposition. Using suitable digital filter

banks, a discrete time signal is decomposed into orthogonal

components, which also have disjoint frequency changes.

Then, by combining various orthogonal components one can

develop representations of the signal at various resolution

levels. For cases of nonstationary ecological signals, such

as the Eh time series realized in the tidal microcosms of the

current study, periodic features, transients, and long-term

trends can be determined without unknown aliasing in

resulting spectra. A further advantage is that the effects of

low-resolution sampling can be examined, and the resulting

effects of aliasing on the signal analysis evaluated. This also

was applied to the Eh time series as an illustration of the

perils of undersampling outputs dynamic systems and then

applying spectral analytical or other analytical techniques to

those data.

Synthetic Organic Chemistry

In order to quantify minor or subtle differences in

transformation of target compounds between treatments, it is

desirable to have stable isotope-labeled, e.g., deuterated,

standards for each target compound of interest. These

standards can be added to sediments before extraction and

serve as internal “monitors” of extraction and analytical

efficiencies for the target compounds. As these deuterated

standards have properties virtually identical to the target

compounds, their loss in the various steps of extraction and

analysis is the same as the compounds under study. With

appropriate controls and calibration, this allows for

quantitative analyses (percent differences can be detected),

rather than semiquantitative data (order-of-magnitude

differences are detected) provided by many analytical

approaches (Catallo 1996a). In addition to this advantage,

isotopic dilution also allows for data streamlining:

quantitation is simply a matter of integral ratios between the

standard (known concentration and base peak integral) and

the target analyte (concentration unknown, base peak

integral known). Unfortunately, commercial availability of

deuterated NOSHs is very limited. As a result, methods were

developed in our laboratory to label the target compounds

with deuterium either using de novo or postsynthetic

approaches (reviewed in Junk and Catallo 1996, 1997), the

latter typically using supercritical (T >374 oC/221 bar)

deuterium oxide (D2O), or “heavy water.” A basic example of

this strategy is shown for the N-heterocycle, quinoline,

below:

n
34

Supercritical water is extremely corrosive, and commercial

stainless steel reactors leaked or exploded (destroying

furnaces and part of a wall) after a few hours. As a result, a

substantial amount of effort in the first year of this work was

devoted to designing reactors that could withstand

supercritical aqueous conditions and were large enough to

permit multigram labeling of target compounds. With this

accomplished, the target compounds were made in

quantities sufficient for use in the transformation studies of

NOSHs in the microcosms. Unfortunately, some NOSHs

were unstable even in subcritical water, and novel synthetic

protocols had to be devised for these compounds. Thus,

quinoxaline-d6, 2-methylquinoxaline-d8, and 2,3-

dimethylquinoxaline-d10 also were prepared by syntheses

developed for this work (Junk and Catallo 1997, Junk and

others 1997). Attempts to prepare phenazine-d8 and

acridine-d9 by: (a) base-catalyzed supercritical isotope

exchange at 400 o

C for 6 hours following Junk and Catallo

(1996); (b) acid catalyzed isotope exchange at 220 oC for 2

days; and (c) palladium-catalyzed exchange at 250 oC for 2

days, all resulted in extensive substrate decomposition. All

compounds subsequently were prepared by base-catalyzed

near-critical exchange at 300 oC for 4 hours. Exchange was

carried out by heating 300 mg substrate, 15 mL D2O, and 0.1

mL 40 percent NaOD (in D2O) in a 30 mL Hastelloy C-22

autoclave. The crude products were extracted with

dichloromethane (DCM), the solvent evaporated and

deuterated quinoxaline, and 2,3-dimethylquinoxaline purified

with a microdistillation apparatus. Phenazine and acridine

were crystallized from methanol. Further purification was

achieved by chromatography (200 mesh silica gel,

hexane:DCM 10:1 v/v). Yields ranged from 45 to 55 percent.

Dibenzothiophene-d8, phenothiazine-d8, phenanthridine-d9

were prepared using the published Supercritical Isotope

Exchange technique (Catallo and Junk 1998, Junk and

Catallo 1996). A 30-mL Hastelloy-C22 autoclave was

charged with the 0.3 g of the respective substrates, 15 mL

deuterium oxide, and 0.1-mL 40-percent deuterium

deuteroxide solution. Exchange was achieved by heating to

400 o

C for 6 hrs. Compounds were collected by filtration and

purified by chromatography over 200-mesh silica gel using

hexane as mobile phase for dibenzothiophene and

phenanthridine, and DCM for phenothiazine. Yields for

dibenzothiophene and phenanthridine were above 75

percent and have been reported along with a range of other

AHs and NOSHs (Junk and Catallo 1996). The yield for

phenothiazine was 86 percent. No attempts were made to

preclude the facile back-exchange of the N-H proton of

phenothiazine during work up under ambient (open-air)

conditions because this analyte would be exposed to water

during the sample extraction.

NOSH Transformation Studies

Three sediment columns prepared as described above were

equilibrated under drained, flooded, and tidal conditions for 2

weeks. The sediments then were collected from the surface

third of the enclosures from each hydrologic type. Protiated

(hydrogenated) NOSH compounds in acetone were added to

the sediments with mechanical mixing (2 hours) to provide

uniform levels of the individual target contaminants between

200 and 400 parts per million (ppm) on a weight basis. Care

was taken to maintain the electrochemical condition of the

sediments (oxidized vs. reduced) during mixing by purging

the system and flooding the headspace of the mixing

enclosure with Ar (reduced, flooded) or air (oxidized,

drained, and tidal). The sediments then were reintroduced to

their respective microcosms in the appropriate column

enclosure. The three hydrologic regimes were initiated and

“time zero” samples were collected within 8 hours of

placement using a clean glass corer. Subsequent samples

were taken at intervals (1 to 4 weeks) depending on

estimators of microbial activity, e.g., iodonitrotetrazolium

reduction (Catallo and others 1990), and previous recovery

of NOSH target analytes. After collection, the sedimentwater samples (ca. 10 g) were weighed and amended with

deuterated isotopic dilution standards for each NOSH target

analyte at the levels near the sediment concentrations. The

samples then were Soxhlet extracted with DCM for 48 hours,

with the extract subsequently dried (Na2

SO4) and

concentrated under N2. The residual sediments in the

extraction thimbles were dried and weighed. The sample

extracts then were subjected to GC-MS in the full-scan

mode. NOSH target analyte concentrations in the extract

were determined vs. the isotopic dilution standard by ratios

of corresponding peak integrals. These extract

concentrations were corrected for concentration factor and

dry weight of sediment in the original sample.

RESULTS AND DISCUSSION

The biogeochemical behavior of the sediment columns in

the microcosms was very similar to that observed in natural

settings and in accord with prevailing theory (Catallo 1999,

Jorgensen and Okholm-Hansen 1985, Odum 1981). With

respect to the Eh, the continuously flooded and drained

systems had reduced (mean = -428 mV) and oxidized (+ 73

mV) Eh values, respectively, with no evidence of daily or

longer periodic variation. The routine (weekly) wetting/

draining of the drained system, and water exchange in the

flooded system, did not affect these trends; apparently the

pre-equilibration of sediments prior to addition of the NOSH

analytes afforded a degree of redox potential poising, i.e.,

“buffering,” that was not affected by the brief maintenance

interventions. The tidal systems, however, exhibited

oscillating Eh values, with significant amplitudes (40 to 180

mV) (fig. 3). Analysis of these signals using the wavelet

transform confirmed the presence of strong diurnal Eh

variations, with a mean value period of 23.78 ± 2.10 hours. It

has been shown that these diurnal signals were: (1)

reproducible in the tide microcosms; (2) found in the

corresponding tidal mesocosms, which also contained the

plant S. alterniflora; and (3) observable in tidal field sites but

not impounded areas isolated from diurnal tides (Catallo

1999). Thermal and light cycles in the laboratory had no

discernable effect on the Eh time signature; variation of tide

stage throughout the day and at different times of the year

N

H H

H

H

H

H

H

D O2

ODN

D

D

D

D

D

D

D
35

(the experiments were conducted in an annex not equipped

with climate control) did not affect the tide-derived Eh signal

period.

The Eh traces shown in figure 3 are characteristic of the tidal

systems in three main respects: (1) the sinusoidal Eh

oscillations continued for as long as the tides were applied,

but ceased immediately when static conditions ensued; (2)

the waveforms are asymmetric, exhibiting hysteresis with

respect to oxidation and reduction; and (3) electrodes at the

different depths were out of phase (as expected), with

deeper electrodes showing smaller amplitudes (probably

reflecting the effects of compaction). These features were

consistent throughout the experimental runs and in

subsequent work with large hydrodynamic mesocosms

containing plants and invertebrates (Catallo 1999). The

hysteresis of measured redox potentials (item 2, preceding)

suggests that the Pt electrode-SCE cell is quasi-reversible

with respect to oxidation and reduction in sediments; i.e., it is

patently non-Nernstian. This almost certainly reflects

compositional as well as bulk phase chemical variables in

situ. Unpacking these factors and adequately describing the

physics involved is far beyond the scope of this

communication and might well be impossible given the

current level of theory on heterogeneous electrochemical

systems (Southampton Electrochemistry Group 1985). In

spite of this, the electrodes in the tidal systems remained

functional throughout the duration of the experiments and

rarely had to be replaced. The electrodes in the oxidized and

reduced static microcosms, however, were replaced

frequently because of passivation with oxides and sulfides

(respectively) and consequent loss of calibration. It would

seem that dynamic Eh conditions in the tidal systems

reduced the level of passivation on the working electrodes

and provided for conditions most favorable for obtaining

accurate potential values vs. the more static, well-poised

conditions in the other treatments. “Reconditioning” of solid

electrode surfaces by alternating oxidation-reduction cycles,

i.e., cyclic removal of surface metal oxides and sulfides by

alternating redox processes, apparently has not been

reported elsewhere and represents an area of interest for

further research.

Biogeochemical gas mixtures leaving the sediments also

reflected the hydrological status of the treatments: the major

gases observed were CO2 from microbial respiration

(drained and tidal) vs. large amounts of sulfide and methane

(flooded conditions) (fig. 4). The gas profiles in each

treatment were checked weekly and were consistent

throughout the experiment.

Transformation of the NOSH analytes was significantly

different in each of the hydrologic regimes, with the rate of

transformation and the number of NOSHs degraded

decreasing in the following order: drained > tidal > flooded.

This can be discerned in figures 5 through 7 with respect to

the number of NOSHs transformed so as to afford a

recovery of under 50 ppm at 16 weeks. Except for one

sample showing a spike at week 18, the drained system

showed the most complete degradation of quinoxaline, 2-

Figure 3—Time series of Eh from shallow and deep electrodes placed in tide simulation microcosms.

Time (days)
36

Figure 4—Headspace gases from the tidal (A) and flooded (B) microcosms by FTIR. (Gases from the drained system were similar to A.)

Water

Carbon dioxide

(A)

(B)

Methane and

mixed sulfides

Carbon dioxide Methane

Wave number (per cm)

Transmittance Transmittance

Figure 5—NOSH transformation profiles from the drained/

oxidized microcosm.

Quinoxaline

2-methylquinoxaline

2,3-dimethylquinox

Phenazine

Dibenzothiophene

Acridine

Phenanthridine

Phenothiazine

Quinoxaline

2-methylquinoxaline

2,3-dimethylquinox

Phenazine

Dibenzothiophene

Acridine

Phenanthridine

Phenothiazine

Figure 6—NOSH transformation profiles from the diurnal tidal

microcosm.
37

methylquinoxaline, 2,3-dimenthylquinoxaline, and

phenanthridine by week 16, with the recovered

concentrations of the other NOSHs also diminished to or

below 25 ppm, i.e., a tenfold concentration reduction. Under

tidal conditions, quinoxaline, 2-methylquinoxaline, and

phenothiazine were eliminated more rapidly than in any

other treatment, with phenazine and phenanthridine reduced

below 25 ppm by week 16. By 17.5 weeks, all NOSH

analytes were transformed so that recoveries were less than

25 ppm. Under flooded conditions, only quinoxaline and

dimethylquinoxaline were degraded to recovery levels below

50 ppm within 16 weeks, and these were completely

eliminated by week 17. Other than these, only phenothiazine

and 2,3-dimethylquinoxaline were transformed to below 50

ppm recovery by week 20. The sediment concentrations of

the other NOSHs remained static, with recovery levels

remaining relatively constant between 100 and 170 ppm for

acridine, phenanthridine, and phenazine between weeks 7

and 20. It would seem then, that except for quinoxaline, 2-

methylquinoxaline, and perhaps phenothiazine, there was a

pronounced Eh effect on transformation rates for the NOSH

compounds studied here.

These transformation studies, while preliminary, clearly

indicated that tidal pulsing optimized the transformation of

some NOSH compounds relative to flooded conditions, and

that in situ or “passive” remediation of coal- and petrochemical pollution in coastal wetlands should include design

features that accommodate prevailing hydroperiods,

including tides and seasonal events. For most settings

where NOSH and AH contamination is a problem, the

statically well-drained/oxidized approach is not an option

unless excavation or draining/impounding of the system is

envisaged. In these cases, changes in sediment

physicochemical properties, e.g., acidification, resulting from

such approaches are not easily reversible and would

attenuate the usability of these sediments in post-treatment

applications, e.g., marsh restoration. On the other hand, in

many cases so called passive remediation approaches

could be adopted that exploit natural hydroperiodicities and

the biogeochemical processes that are coupled to them,

rather than artificially manipulating them, i.e., by impounding

or dredging, and further compromising the functioning of the

ecosystem.

In an abstract sense, the use of hydroperiodicity for

enhanced chemical remediation and ecological recovery of

polluted systems seems promising in light of this and related

work (Catallo 1996b). It would involve, at the least, attempts

to optimize the tidal volume of the wetland without

compromising its integrity and function. Obviously, this kind

of undertaking in a real wetland would involve a set of

engineering interventions and monitoring strategies that

encompass hydrologic, sedimentologic, and biogeochemical

variables in a progressive sense. The same is true of the use

of marginally contaminated dredge materials and other

wastes, e.g., phosphogypsum, bauxite red mud, for coastal

habitat restoration projects. The author is aware of no actual

cases in which the goals and approaches of chemical

remediation and wetland creation/restoration have been

successfully coupled. Much further ecological study is called

for in microcosms and other controlled settings where

variables and causal relationships can be identified and

ranked with respect to holistic endpoints including, but not

limited to, pollutant transformation.

ACKNOWLEDGMENTS

The author is indebted to Professors R.D. DeLaune, R.J.

Gale, R.P. Gambrell, J. Aravena, T. Junk, and R. Meyhandan

of LSU for discussion, critique, and substantive assistance.

R.S. Murray, L.D. Civils, N. Crochet, and H.L. Dharmasena

provided assistance with field work, laboratory, and

greenhouse studies, logger development, and synthetic

chemistry. This work was supported in part by grants from

USDOI, USEPA, and USFWS.

REFERENCES

Asher, H.B. 1983. Causal modeling. SAGE Univ. Pap. 3. Newbury

Park, CA: Sage Publishers. 91 p.

Brockwell, P.J.; Davis, R.J. 1991. Time series: theory and methods.

New York: Springer Verlag. 577 p.

Catallo, W.J. 1993. Ecotoxicology of wetland ecosystems. Current

understanding and future research needs. Annual review.

Environmental Toxicology and Chemistry. 12(12): 2209–2224.

Catallo, W.J. 1996a. Bioremediation: statistical and analytical needs.

In: Proffitt, C.E.; Roscigno, P.F., eds. Proceedings: Gulf of Mexico

and Caribbean oil spills in coastal ecosystems: assessing

effects, natural recovery, and progress in remediation research.

OCS Study/MMS 95–0063. New Orleans: U.S. Department of the

Interior, Mineral Management Service, Gulf of Mexico Region:

109–129.

Catallo, W.J. 1996b. Transformation of N-, O-, and S-heterocycles

(NOSHs) in estuarine sediments: effects of redox potential and

particle size. Chemosphere. 33/12: 2543–2563.

Catallo, W.J. 1999. Hourly and daily variation of sediment redox

potential in tidal wetland sediments. Biol. Sci. Rep.: USGS/BRD/

BSR–01–1999. U.S. Geological Service, Biological Resources

Division.

Catallo, W.J.; Gale, R.J.; Wong, R.L.; Bender, M.E. 1990. The

reducible dye 2-(p-iodophenyl)-3-(p-nitrophenyl)-5-phenyl-2Htetrazolium chloride (INT) for use in aquatic toxicology. Notes on

chemical structure, electrochemistry, and toxicity. In: Landis,

Figure 7—NOSH transformation profiles from the flooded/reduced

microcosm.

Quinoxaline

2-methylquinoxaline

2,3-dimethylquinox

Phenazine

Dibenzothiophene

Acridine

Phenanthridine

Phenothiazine
38

W.G.; van der Schalie, W.H., eds. Aquatic toxicology and risk

assessment. ASTM Spec. Tech. Publ. 1096. Philadelphia:

American Society for Testing and Materials: 222–236.

Catallo, W.J.; Gambrell, R.P. 1994. Fates and effects of N-, O-, and

S-heterocycles (NOSHs) from petroleum and pyrogenic sources

in marine sediments. Final Grant Completion Rep. OCS Study/

MMS 674–2010. New Orleans: U.S. Department of the Interior,

Minerals Management Service, Gulf of Mexico, OCS Regional

Office. 72 p.

Catallo, W.J.; Gambrell, R.P.; Reddy, K.R. [and others]. 1999.

Biogeochemical processes. In: Lewis, M.A.; Mayer, F.L.; Powell,

R.L. [and others], eds. Ecotoxicology and risk assessment for

wetlands. Pensacola, FL: SETAC Foundation for Environmental

Education: 27–69.

Catallo, W.J.; Hoover, D.H.; Vargas, D. 1994. Effects of quinoline on

ultrastructure in Paramecium caudatum and the protective

effects of dissolved calcium. Aquatic Toxicology. 29(324): 291–

303.

Catallo, W.J; Junk, T. 1998. Process for preparing deuterium tagged

compounds. U.S. patent 5,830,763. Nov. 3.

Jorgenesen, B.B.; Okholm-Hansen, B. 1985. Emissions of biogenic

sulfur gases from a Danish estuary. Atmospheric Environment.

19(11): 1737–1749.

Junk, T.; Catallo, W.J. 1996. Preparative supercritical deuterium

exchange in arenes and heterocycles. Tetrahedron Letters. 37:

3445.

Junk, T.; Catallo, W.J. 1997. Organic reactions in supercritical polar

fluids. Royal Society of Chemists (U.K.). Chemical Society

Reviews. 26: 401–406.

Junk, T., Catallo, W.J.; Elguero, J. 1997. Synthesis in superheated

aqueous media: preparation of fully deuterated pyrazoles and

quinoxalines. Tetrahedron Letters. 38(36): 6309–6312.

Lindsay, W.L. 1991. Iron oxide solubilization by organic matter and

its effect on iron availability. In: Chen, Y.; Haddar, Y., eds. Iron

nutrition and interactions in plants. The Netherlands: Kluwer

Academic Publishers: 29–36.

Mallat, S. 1998. A wavelet tour of signal processing. New York:

Academic Press. 577 p.

Odum, H.T. 1983. Ecological and general systems. Boulder, CO:

University of Colorado Press. 644 p.

Picek, T.; Simek, M.; Santruckova, H. 2000. Microbial responses to

fluctuation of soil aeration status and redox conditions. Biology

and Fertility of Soils. 31: 315–322.

Southampton Electrochemistry Group. 1985. Instrumental methods

in electrochemistry. Chichester, West Sussex, UK: Ellis Horwood

Ltd. 443 p.

Turov, Y.P.; Gerasimova, N.N.; Sagschenko, T.A.; Bieko, O.A. 1987.

Group composition of the low molecular weight nitrogen bases of

Samotlor crude oil. Petroleum Chemistry USSR. 27(1): 20–25.

Warshawsky, D. 1992. Environmental sources, carcinogenicity,

mutagenicity, metabolism, and DNA binding of nitrogen and

sulfur heterocyclic aromatics. Environ. Carcino. Ecotox. Revs.

C(10): 1–71.
39

INTRODUCTION

Deepwater swamps are found along rivers and streams of

the Atlantic and Gulf Coastal Plains, and baldcypress

[Taxodium distichum (L.) Rich.], pondcypress (T. distichum

var. nutans), water tupelo (Nyssa aquatica L.), and Atlantic

white-cedar (Chamaecyparis thyoides) are common tree

species found in these areas. Both the Spanish in Florida

and French in Louisiana found Indians using cypress (when

used this way can be either baldcypress or pondcypress),

which the Seminoles called “hatch-in-e-haw,” meaning

everlasting (Neubreck 1939). Europeans quickly recognized

that cypress wood was very rot resistant, strong, and easily

worked, and efforts to establish a timber trade with

Louisiana began around 1700 (Mancil 1980). Cypress was

the staple commodity of the colonial lumber industry in

Louisiana and the principal cash product for most colonists

of the Lower Mississippi Valley until the 1790s, when sugar

products became profitable.

The timber resource in the swamps seemed inexhaustible to

early settlers with over 35.5 million m3

 (15 billion fbm) of

cypress estimated in the Louisiana delta swamps alone

(Kerr 1981). Harvesting in these wet swamps was seasonal

in nature until the invention of the pullboat in 1889.

Pullboats and the expansion of the railroad system

(Sternitzke 1972), combined with a massive national

campaign by cypress dealers (Burns 1980), resulted in a

logging boom during the period from 1890 to 1925.

Production of cypress lumber increased from 1.17 million m3

(495 million fbm) in 1899 to over 2.36 million m3 (1 billion

fbm) in 1913 (Betts 1938, Mattoon 1915). By 1925, nearly all

of the virgin timber had been cut and most of the mills

closed. In 1933, only about 10 percent of the original

standing stock of cypress remained (Brandt and Ewel

1989), but some cypress harvesting continued throughout

the Southern United States on a smaller scale (Conner

1988).

Atlantic white-cedar logging began as early as 1700 in North

and South Carolina (Frost 1987) and 1749 in New Jersey

(Little 1950). Up to 50 percent of the Atlantic white-cedar

area in North Carolina was cut between 1870 and 1890. As

in other parts of the Southern United States, the rate at

which Atlantic white-cedar swamps were logged greatly

increased following the introduction of railroads, steam

logging technology, portable sawmills, and dredging

technology (Earley 1987, Frost 1987).

Unfortunately, the early exploitation of these swamplands

occurred with little regard for sustainability. According to one

logger, “We just use the old method of going in and cutting

down the swamp and tearing it up and bringing the cypress

out. When a man’s in here with all the heavy equipment, he

might as well cut everything he can make a board foot out of;

we’re not ever coming back in here again” (Van Holmes

1954). Nearly all of the virgin swamplands in the Southern

United States were logged (Conner and Buford 1998).

Compared to other forest ecosystems, little silvicultural

information is available for deepwater swamp forests, and

management of these areas has been largely limited to

clearcutting and highgrading (Johnson 1979, Williston and

others 1980). Only recently have studies begun to

investigate the response and recovery of these forests to

harvesting practices, e.g., Aust and others 1989, 1997.

Cypress and water tupelo regenerate well in swamps where

the seedbed is moist and competitors are unable to cope

with flooding, but extended dry periods are necessary for the

seedlings to grow tall enough to survive future flooding

(Keeland and Conner 1999). Early height growth is important

because seedlings can be killed in as little as 10 to 12 days

of total submergence during the growing season (Demaree

1932). Although natural regeneration has also been the

preferred method of regenerating Atlantic white-cedar

RESTORATION METHODS FOR DEEPWATER SWAMPS

William H. Conner, Kenneth W. McLeod, and Ellen Colodney1

1 Connor, Professor, Baruch Institute of Coastal Ecology and Forest Science, Clemson University, Box 596, Georgetown SC 29442; McLeod,

Associate Professor, Savannah River Ecology Laboratory, Drawer E, Aiken, SC 29801; Colodney, Owner, Coastal Plain Conservation Nursery,

3067 Conners Drive, Edenton, NC 27932.

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability of

wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.

Abstract—Planting in deepwater swamp areas is difficult and time consuming, and nursery-grown seedlings are often not

suited for such conditions. Baldcypress [Taxodium distichum (L.) Rich.], water tupelo (Nyssa aquatica L.), swamp blackgum

[N. sylvatica var. biflora (Walt.) Sarg.], and green ash (Fraxinus pennsylvanica Marsh.) have been planted at various

flooded sites in South Carolina and Louisiana. One of the most effective means of planting these species in flooded

situations was to heavily prune the lateral roots, grasp the seedling at the root collar, and push it into the soil. Excellent

results have been obtained with baldcypress, whereas green ash was most sensitive to root pruning and water depth. Water

tupelo and swamp blackgum were intermediate in response. Tree shelters are commonly used to reduce herbivory problems, and height growth inside the shelters is increased. Additional research is needed to compare operational performance

of various techniques under conditions of interacting stresses such as herbivory and flooding.
40

(Laderman 1989), many projects now rely on artificial means

to recreate Atlantic white-cedar habitat (Guidry 1999,

Phillips and others 1993), and much research has been

initiated to identify nursery practices to produce the best

seedlings (Summerville and others 1999). Lack of seed

source and herbivory problems are commonly reported as

causing failure of natural regeneration projects for Atlantic

white-cedar.

Coppice regeneration is also a possibility in cutover areas of

cypress and water tupelo. Stumps of vigorous stock up to 60

years old can generally be counted on to send up healthy

sprouts. Although many stumps sprout during the first

growing season after logging, few of these sprouts survive,

although results are often contradictory.

PROBLEMS LIMITING SWAMP REGENERATION

Flooding and Salinity

Human activities have inextricably altered the hydrology of

almost all major alluvial floodplains in the United States

within the last two centuries through the construction of

dams, levees, and causeways and by channelization. Dams

reduce the frequency and magnitude of downstream

flooding, often extend the length of time the floodplain is

inundated, and reduce the rates of erosion and

sedimentation. Channelization and canal building, with

associated levees or spoil banks, represent major

modifications to natural hydrological patterns and often

result in permanent impoundment of large areas of

swamplands (Conner and Day 1989). Floodplain

communities are adapted to a fairly predictable flood pulse,

and alteration of the timing, duration, or magnitude of this

flooding reduces diversity and productivity (Junk and others

1989). Because many swamp areas are permanently to

nearly permanently flooded, natural regeneration is

negligible (Conner and others 1986), and planting is difficult.

Another aspect of flooding that should be considered for

vast coastal swamp areas is sea level rise and resulting

increases in salinity (Conner and Brody 1989, Conner and

Day 1988). While baldcypress and water tupelo can survive

extended and even deep flooding (Hook 1984, Keeland and

Sharitz 1995), they do not seem capable of enduring

sustained flooding by water with salinity levels > 8 ppt

(Conner and others 1997, McLeod and others 1996), and

Atlantic white-cedar is very intolerant of salinity (Little 1950).

Hurricanes

Coastal Plain swamps have developed with windstorms as a

normal, episodic part of the climatic regime (Conner and

others 1989). Recent hurricanes such as Hugo (1989) in the

southeast Atlantic Coastal Plain and Andrew (1992) in the

northern Gulf caused extensive damage to forests in their

paths. Such damage may be especially severe to the

shallow-rooted hardwoods with large crowns that are

common on alluvial floodplains. In the Congaree Swamp in

South Carolina, 61 percent of the bottomland oaks (Quercus

spp.) and 45 percent of the sweetgums were severely

damaged by Hurricane Hugo, but only 3 percent of the

baldcypress trees were affected (Sharitz and others 1993).

Regeneration in hurricane-damaged areas may be limited if

natural hydrological patterns have been altered.

Animals

Nutria were introduced from South America in the 1930s,

and early plantings of baldcypress in Louisiana were

destroyed (Blair and Langlinais 1960). The problem has not

been solved (Brantley and Platt 1992, Conner 1988), and

nutria have been reported to damage even mature trees

(Hesse and others 1996). Beaver, deer, and feral hogs can

also present a problem and can be fairly numerous in some

areas. Conner and others (2000) found that clipped

seedlings usually die, but baldcypress tends to resprout in

many cases. Even such a small creature as the crayfish can

become a problem to planted seedlings when food sources

are low. Scraping of algae at the waterline can girdle a

seedling and cause tip die-back (Conner 1988). Deer, field

mice, and rabbits have deterimental effects on Atlantic

white-cedar seedlings (Guidry 1999, Zimmerman and others

1999).

RESTORATION OF WETLAND SITES

Because of loose, unconsolidated muck commonly found in

deepwater swamps, an easier method of planting seedlings

was needed. The method adopted in our studies involves

pruning the lateral root systems and clipping the tap root so

we end up with a spear. Root systems of seedlings grown in

unsaturated soils in the nursery are not appropriate to

saturated soils, and large portions of this system will be lost

once planted. A new root system appropriate for saturated

soils will be produced. Since much of the original root

system would normally be lost anyway, pruning it prior to

planting generally does not cause problems.

Seedlings are bundled in plastic with damp peat moss

around the tap root and either stored in coolers or

transported to the field. Planting can be done along precisely

laid out lines or by walking and estimating distances. By

holding the seedling at the root collar, one pushes the

seedling into the ground until the hand hits the soil surface.

There are no tools to carry for digging holes, and one does

not have to worry about filling in completely around the root.

In very loose soils, the seedlings will need to be staked to

keep vertical.

Root pruning does not work well with all species. While

baldcypress and tupelo success rates have been high,

green ash (Fraxinus pennsylvanica Marsh.) and swamp

blackgum [N. sylvatica var. biflora (Walt.) Sarg.] success has

been poor. Green ash seemed to do well in the first 1 to 2

years after planting but died in succeeding years (Conner

and others 2000). One reason is that root systems never

redeveloped on the seedlings when planted in wet areas.

Hand-bagged seedlings and balled and burlap seedlings

have also been tried (Conner and others 1999). Balled

seedlings planted on the sediment surface produced

sufficient rooting down into the sediment to withstand

complete drying of the surface water. However, there was no

real benefit to using the hand-bagged or balled and

burlapped seedlings since root-pruned baldcypress and

water tupelo seedlings are less costly and easier to plant

and survive just as well.

If the site is subject to drying out, a seedling with a large

lateral root component is desirable. Increased root branching
41

allows for more water and nutrient absorption, which results

in more root growth and a healthier seedling. One such way

to accomplish this is by way of specialized containers such

as RootMaker. The pots have openings in the corners and at

various levels in the sides and bottom. Roots are air-pruned

and the design is such to prevent the secondary roots from

becoming congested at the bottom of the pot. Seedlings are

generally grown on a wire bench 20 cm above the moist soil

to allow for good air circulation. Otherwise the roots of

species like baldcypress and some oaks grow out of the

openings, swell, and removal is awkward. The roots and

secondary roots are well distributed within the pot.

Atlantic white-cedar is generally done using the same

planting tools and techniques used in planting southern

pines, and there is no apparent need to develop other

techniques. Work by Weyerhauser has shown that rooted

cuttings are more reliable and easier to grow than bare-root

seedlings (Phillips and others 1998).

Tree shelters work well, especially in early years before

seedlings emerge from the tube. After emerging from

shelters, height growth is not as great, but the initial growth

spurt keeps them above nonsheltered seedlings. Overall,

baldcypress and water tupelo have grown well (Conner and

others 2000). Tree shelters are generally advertised to

breakdown within a few years due to ultraviolet light in the

field. This has not been our experience, but the diameter

growth of the seedlings has been strong enough to rip the

shelters longitudinally and not girdle the trees.

CONCLUSIONS

Tree species composition in deepwater swamps remains

fairly constant because so few species can tolerate

extended flooding. With changes in hydrology, many of these

forests are now flooded more than in the past. Regenerating

swamp forests is not a simple matter of overcoming past

disturbances, but is complicated by continuing disturbances

and impacts, both natural and manmade. Successful

regeneration is limited by flooding, and planting may be

required to ensure that populations of these trees are

established in areas impacted by manmade (logging) or

natural (herbivory, hurricanes, salt-water intrusion)

disturbances. Not all adverse site conditions can be

overcome, and some solutions may not be cost effective in

meeting landowner objectives. Simple root-pruning

techniques allow easy planting of these areas, and survival

and growth have been shown to be excellent in many areas

if animal herbivory can be controlled.

ACKNOWLEDGMENTS

Many field volunteers including L. Wayne Inabinette, Michael

R. Reed, and Keith T. Barnett have aided the research that

has been conducted through the years on regeneration.

Funding for this research has come from a variety of

sources including McIntire-Stennis funds from Clemson

University, the Savannah River Technology Center

(Westinghouse Savannah River Company), and Financial

Assistance Award Number DE–FC09–96SR18546 from the

Department of Energy to the University of Georgia Research

Foundation.

REFERENCES

Aust, W.M.; Mader, S.F.; Lea, R. 1989. Abiotic changes of a tupelocypress swamp following helicopter and rubber-tired skidder

timber harvest. In: Miller, J.H., comp. Proceedings of the fifth

biennial southern silvicultural conference; 1988 November 1–3;

Memphis, TN. Gen. Tech. Rep. SO–74. New Orleans: U.S.

Department of Agriculture, Forest Service, Southern Forest

Experiment Station: 545–551.

Aust, W.M.; Schoenholtz, S.H.; Zaebst, T.W.; Szabo, B.A. 1997.

Recovery status of a tupelo-cypress wetland seven years after

disturbance. Forest Ecology and Management. 90: 161–169.

Betts, H.S. 1938. Southern cypress. American woods. Washington,

DC: U.S. Department of Agriculture, Forest Service, U.S.

Government Printing Office. 6 p.

Blair, R.M.; Langlinais, M.J. 1960. Nutria and swamp rabbits damage

baldcypress plantings. Journal of Forestry. 58: 388–389.

Brandt, K.; Ewel, K.C. 1989. Ecology and management of cypress

swamps: a review. Gainesville, FL: Florida Cooperative Extension Service, Institute of Food and Agricultural Sciences,

University of Florida.

Brantley, C.G.; Platt, S.G. 1992. Experimental evaluation of nutria

herbivory on baldcypress. Proceedings of the Louisiana

Academy of Science. 55: 21–25.

Burns, A.C. 1980. Frank B. Williams: cypress lumber king. Journal of

Forest History. 24: 127–133.

Conner, W.H. 1988. Natural and artificial regeneration of

baldcypress in the Barataria and Lake Verret Basins of Louisiana. Baton Rouge, LA: Louisiana State University. 148 p. Ph.D.

dissertation.

Conner, W.H.; Brody, M. 1989. Rising water levels and the future of

southeastern Louisiana swamp forests. Estuaries. 12: 318–323.

Conner, W.H.; Buford, M. 1998. Southern deepwater swamps. In:

Messina, M.G.; Conner, W.H., eds. Southern forested wetlands:

ecology and management. Boca Raton, FL: Lewis Publishers/

CRC Press: 261–287.

Conner, W.H.; Day, J.W., Jr. 1988. Rising water levels in coastal

Louisiana: implications for two forested wetland areas in

Louisiana. Journal of Coastal Research. 4: 589–596.

Conner, W.H.; Day, J.W., Jr. 1989. Response of coastal wetland

forests to man-made and natural changes in the environment

with emphasis on hydrology. In: Hook. D.D.; Lea, R., eds.

Proceedings of the symposium: forested wetlands of the

Southern United States. Gen. Tech. Rep. SE–50. Asheville, NC:

U.S. Department of Agriculture, Forest Service, Southeastern

Forest Experiment Station: 34–43.

Conner, W.H.; Day, J.W., Jr.; Baumann, R.H.; Randall, J. 1989.

Influence of hurricanes on coastal ecosystems along the

northern Gulf of Mexico. Wetlands Ecology and Management. 1:

45–56.

Conner, W.H.; Inabinette, L.W.; Funderburke, E.L. 2000. The use of

tree shelters in restoring forest species to floodplain delta: 5-year

results. Ecological Engineering. 15(Suppl. 1): S47–S56.
42

Conner, W.H.; McLeod, K.W.; Inabinette, L.W. [and others]. 1999.

Successful planting of tree seedlings in wet areas. In: Haywood,

J.D., ed. Proceedings of the tenth biennial southern silvicultural

research conference. Gen. Tech. Rep. SRS–30. Asheville, NC:

U.S. Department of Agriculture, Forest Service, Southern

Research Station: 201–204.

Conner, W.H.; McLeod, K.W.; McCarron, J.K. 1997. Flooding and

salinity effects on growth and survival of four common forested

wetland species. Wetlands Ecology and Management. 5: 99–

109.

Conner, W.H.; Toliver, J.R.; Sklar, F.H. 1986. Natural regeneration of

cypress in a Louisiana swamp. Forest Ecology and Management. 14: 305–317.

Demaree, D. 1932. Submerging experiments with Taxodium.

Ecology. 13: 258–262.

Earley, L.S. 1987. Twilight for junipers. Wildlife in North Carolina. 51:

9–15.

Frost, C.C. 1987. Historical overview of Atlantic white cedar in the

Carolinas. In: Laderman, A.D., ed. Atlantic white cedar wetlands.

Boulder, CO: Westview Press: 257–264.

Guidry, J.I. 1999. Effects of herbivore pressure on Atlantic whitecedar rooted-cutting survival on a North Carolina coastal plain

peatland. In: Shear, T.H.; Summerville, K.O., eds. Proceedings:

Atlantic white-cedar: ecology and management symposium;

1997 August 6–7; Newport News, VA. Gen. Tech. Rep. SRS–27.

Asheville, NC: U.S. Department of Agriculture, Forest Service,

Southern Research Station: 18–21.

Hesse, I.D.; Conner, W.H.; Day, J.W, Jr. 1996. Herbivory impacts on

the regeneration of forested wetlands. In: Flynn, K.M., ed.

Proceedings of the southern forested wetlands ecology and

management conference; 1996 March 25–27; Clemson, SC.

Clemson, SC: Clemson University: 23–28.

Hook, D.D. 1984. Waterlogging tolerance of lowland tree species of

the South. Southern Journal of Applied Forestry. 8: 136–149.

Johnson, R.L. 1979. Timber harvests from wetlands. In: Greeson,

P.E.; Clark, J.R.; Clark, J.E., eds. Wetland functions and values:

the state of our understanding. Minneapolis: American Water

Resources Association: 598–605.

Junk, W.J.; Bailey, P.B.; Sparks, R.E. 1989. The flood-pulse concept

in river floodplain systems. Canadian Special Publication of

Fisheries and Aquatic Sciences. 106: 110–127.

Keeland, B.D.; Conner, W.H. 1999. Natural regeneration and growth

of Taxodium distichum (L.) Rich. in Lake Chicot, Louisiana after

44 years of flooding. Wetlands. 19(1): 149–155.

Keeland, B.D.; Sharitz, R.R. 1995. Seasonal growth patterns of

Nyssa sylvatica var. biflora, Nyssa aquatica, and Taxodium

distichum as affected by hydrologic regime. Canadian Journal of

Forest Research. 25: 1084–1096.

Kerr, E. 1981. The history of forestry in Louisiana. Alexandria, LA:

Louisiana Forestry Association.

Laderman, A.D. 1989. The ecology of the Atlantic white cedar

wetlands: a community profile. Biol. Rep. 85(7.21). Washington,

DC: U.S. Fish and Wildlife Service, Office of Biological Services.

Little, S., Jr. 1950. Ecology and silviculture of whitecedar and

associated hardwoods in southern New Jersey. Bull. 56. New

Haven, CT: Yale University, School of Forestry.

Mancil, E. 1980. Pullboat logging. Journal of Forest History. 24: 135–

141.

Mattoon, W.R. 1915. The southern cypress. Agric. Bull. 272.

Washington, DC: U.S. Department of Agriculture.

McLeod, K.W.; McCarron, J.K.; Conner, W.H. 1996. Effects of

inundation and salinity on photosynthesis and water relations of

four southeastern Coastal Plain forest species. Wetlands

Ecology and Management. 4(1): 31–42.

Neubreck, W.L. 1939. American southern cypress. Trade Promot.

Ser. 194. Washington, DC: U.S. Department of Commerce.

Phillips, R.; Gardner, W.E.; Summerville, K.O. 1993. Plantability of

Atlantic white-cedar rooted cuttings and bare-root seedlings. In:

Brissette, J.C., ed. Proceedings of the seventh biennial southern

silvicultural research conference; 1992 November 17–19; Mobile,

AL. Gen. Tech. Rep. SO–93. New Orleans: U.S. Department of

Agriculture, Forest Service, Southern Forest Experiment Station:

97–104.

Phillips, R.W.; Hughes, J.H.; Buford, M.A. [and others]. 1998. Atlantic

white cedar in North Carolina, U.S.A.: a brief history and current

regeneration efforts. In: Laderman, A.D., ed. Coastally restricted

forests. New York: Oxford University Press, Inc.: 156–170.

Sharitz, R.R.; Vaitkkus, M.R.; Cook, A.E. 1993. Hurricane damage to

an old-growth floodplain forest in the Southeast. In: Brissette,

J.C., ed. Proceedings of the seventh biennial southern silvicultural research conference; 1992 November 17–19; Mobile, AL.

Gen. Tech. Rep. SO–93. New Orleans: U.S. Department of

Agriculture, Forest Service, Southern Forest Experiment Station:

203–210.

Sternitzke, H.S. 1972. Baldcypress: endangered or expanding

species? Economic Botany. 26: 130–134.

Summerville, K.O.; Gardner, W.E.; Hinesley, L.E.; Bardon, R.E.

1999. Atlantic white-cedar plant production. In: Shear, T.H.;

Summerville, K.O., eds. Proceedings: Atlantic white-cedar:

ecology and management symposium; 1997 August 6–7;

Newport News, VA. Gen. Tech. Rep. SRS–27. Asheville, NC: U.S.

Department of Agriculture, Forest Service, Southern Research

Station: 68–75.

Van Holmes, J. 1954. Loggers of the unknown swamp. Saturday

Evening Post. 226: 32–33, 102–105.

Williston, H.L.; Shropshire, F.W.; Balmer, W.E. 1980. Cypress

management: a forgotten opportunity. For. Rep. SA–FR 8.

Atlanta: U.S. Department of Agriculture, Forest Service,

Southeastern Area.

Zimmerman, G.; Mueller, R.; Brown, J. [and others]. 1999. The Penn

Swamp experiments: an overview. In: Shear, T.H.; Summerville,

K.O., eds. Proceedings: Atlantic white-cedar: ecology and

management symposium; 1997 August 6–7; Newport News, VA.

Gen. Tech. Rep. SRS–27. Asheville, NC: U.S. Department of

Agriculture, Forest Service, Southern Research Station: 45–48.
43

WETLANDS IN SOUTHERN THAILAND

Wetlands in Southern Thailand may be classified into nine

categories according to their formation, location, and

morphology.

Open Sea Coasts, Sandy Beaches,

and Offshore Islands

These wetlands are found along the coastlines and offshore

islands of the peninsula. In most cases, they occur in

association with other coastal wetlands, such as mudflats

and mangroves. There are many islands and small rocky

outcrops off the west coast of the peninsula, most of which

have never been surveyed. Some islets are known nesting

sites for the Pacific Reef Egret (Egretta sacra) and terns

such as Stema cougalli, S. sumatrana, S. bergii, and S.

anaethetus. The Beach Thick-Knee (Esacus magnirostris) is

also located, but is restricted to, sand beaches on offshore

islands, whereas both the Malaysian Plover Charadrius

peronii and the Little Tern S. albifrons breed on the sand

beaches of mainland and island coasts.

Intertidal Mudflats and Mangroves

This type of wetlands is most known and is of great

conservation value in Southern Thailand, having enormous

importance in sustaining both inshore capture fisheries and

aquaculture. The most extensive and species-rich mangrove

ecosystems are found along the west coast of the peninsula,

which supported 63 percent of the total mangrove area of

2871 km2 at the end of 1982. There are also several

important mangrove and mudflat sites on the east coast of

the peninsula. Extensive areas have been converted to

shrimp farms.

Intertidal mudflats are important in terms of wildlife

conservation. These wetlands support a huge number of

passage and wintering herons and shorebirds. The

mangroves themselves still support nesting colonies of

cormorants, herons, and a few Lesser Adjutants (Leptoptilos

javanicus), together with considerable numbers of some

birds of prey such as Brahminy Kites (Haliastur indus). Two

other important species are the Brown-winged Kingfisher

(Pelargopsis amauroptera) and the Mangrove Pitta (Pitta

megarhyncha), both of which are restricted to the west coast.

Lower Perennial Rivers

Wetlands of this category in Southern Thailand consist of

meandered rivers and riverine marshes. In most cases,

these areas are deforested, except where narrow fringes of

fresh and brackish water swamp woodlands remain along

the riverbanks. As most riverbanks usually support a high

human population density, relatively few waterfowl are found.

Only certain grounds with strong conservation movement

may support breeding colonies or large roosts of herons or

storks, together with a few Black Kites (Milvus migrans).

Vertical earth banks also support a few nesting Pied

Kingfisher (Ceryle rudis), whereas mud and sand banks

support passage shorebirds as well as the resident River

Lapwing (Vanellus duvaucelii). There are a few colonies of

Plain Sand martins (Riparia paludicola). The rare and local

Jerdon’s Bushchat (Saxicola jerdoni) appears to be

associated with stands of the tall grass (Saccharum

arundinaceum) in riverine floodplains. The species may have

decreased greatly as a result of the burning of such

vegetation in order to open up seasonally inundated alluvial

soils for dry season cultivation.

Large areas of seasonally inundated land also lie along

many rivers in Southern Thailand, particularly along large

rivers such as the Tapi River in Suratthani and the Pak

Phanang River in Nakhonsrithammarat. Many such areas

are utilized for the cultivation of vegetables or rice as the

WETLANDS SYSTEMS IN SOUTHERN THAILAND: THE ESSENTIAL

RESOURCES FOR SUSTAINABLE REGIONAL DEVELOPMENT

Rotchanatch Darnsawasdi and Prasert Chitpong1

1 Darnsawasdi, Faculty of Environmental Management, and Chitpong, Faculty of Natural Resources, Prince of Songkla University, Hatyai,

Songkhla, 90112.

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability of

wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.

Abstract—Parts of Southern Thailand are inundated by water for months annually resulting in various wetlands including,

among others, Tapi River Basin, Pak Panang River Basin, Songkhla Lake Basin, Pangnga Bay, Pattani River Basin, and

Narathiwas Peat Swamp. Most wetlands perform functions such as flood retention, water filtration, bird and wildlife habitat,

and tree growth. These wetlands are invaluable also for the value derived from: (a) commodity products such as timber,

food, and chemicals; (b) nonconsumptive uses such as recreation and tourism; and (c) environmental attributes such as

biodiversity, wildlife, and water quality. These values are derived from their hydrogeologic and biochemical functions, not

dependent on the size of the wetland but on the intrinsic properties of the ecosystem, particularly their location within a

watershed and positioning with respect to rivers, uplands, and seas. As many wetlands have been transformed and

deteriorated, research into these wetland systems and their relationships with human activities are required. Integrated and

participatory approaches to management of these wetlands are also recommended.
44

seasonal flooding recedes, yet they may perhaps be of

considerable value to waterfowl.

Upper Perennial Rivers

Wetlands of this category are usually characterized by large

streams in mountainous or hilly terrain, often in association

with waterfalls and rapids. Because the upper reaches of

many rivers are still largely forested, they may continue to

support a considerable diversity of wildlife, including a great

many nonaquatic species. Some of the important sites are

already enclosed within the boundaries of national parks,

forest reserves, and wildlife sanctuaries.

Several kinds of birds such as Fish-Eagle, Crested

Kingfisher (Ichthyophaga humilis), Sarcogyps calvus, Pavo

muticus, and Megaceryle lugubris are sometimes found in

these wetlands. Many such sites, including many small

streams and torrents, may be of critical conservation

importance for various frogs and toads. There has never

been any comprehensive assessment of the conservation

status of Thai amphibians, however, so that many important

sites remain undocumented.

Freshwater Lakes, Ponds, and Associated Marshes

These wetlands are usually related to the Lower Perennial

Rivers. They comprise abandoned oxbow lakes, inundated

sinkholes, and submerged flood plains. Most sites are small

(less than 500 ha) and are usually surrounded by rice

paddies. A great many lie within 1 to 2 km of major rivers.

Like the Lower Perennial Rivers, almost all have been

utilized mostly for agriculture causing habitat disturbance,

which greatly limits the residence of wildlife. The vegetation

is usually restricted to floating or submerged aquatic plants

alternating with open water and supports such breeding

waterfowl as Pheasant-tailed Jacana (Hydrophasianus

chirurgus). The highest diversity of breeding and wintering

birds is usually found at those sites, which contain extensive

areas of emergent vegetation, especially Phragmites or

Typha. Many such sites are of international or national

importance for their wintering duck populations.

Water Storage Reservoirs

Wetlands of this category are manmade or constructed

wetlands. Most larger reservoirs are constructed for

hydroelectric power generation, and are situated in steep

forested river valleys. They tend to support fewer native

wildlife, but may lead to a new ecosystem that can support

the other wildlife. Many irrigation reservoirs, on the other

hand, are situated in the plains, are relatively shallow, and

show considerable annual fluctuation in area. Such sites

may be of value for wintering and passage wading birds. A

few such sites are known to support wintering

concentrations of ducks as well as such breeding species as

Porphyrio porphyrio.

Rice Paddies

In most circumstances, rice paddies are river flood plains,

which were transformed into rice fields. They constitute an

important and very extensive, seasonally inundated habitat

for birds. The extent to which these areas can be utilized,

however, depends upon the availability of undisturbed

roosting and nesting sites such as clumps of trees and

permanent water bodies. Egrets and herons feed to a

considerable extent in flooded paddies; cormorants utilize

ditches around their margins, and the Asian Open-billed

Stork (Anastomus oscitans) feeds both in flooded and dry

paddies. Areas of hard, dry paddy stubble in the late dry

season are utilized by huge numbers of nesting Oriental

Pratincoles (Glareola maldivarum).

Freshwater Swamp Woodlands

Detailed information on the history of such sites is usually

lacking. In most cases, such swamp woodlands may be

merely the degraded remnants of primary peat swamp

forests. The most disturbed sites, particularly those having

been subject to repeated burning, are species poor and are

usually dominated by Melaleuca leucadendron. Some

Alstonia spathulata may occur in the less disturbed sites, as

at Thale Noi Non-Hunting Area in Peninsular Thailand. Small

areas of other freshwater swamp woodland formations may

occur along the banks of larger rivers and other sites, which

are subject to occasional inundation. Such sites are of

considerable importance for nesting and roosting colonies of

larger waterbirds, such as cormorants, herons, and storks.

Peat Swamp Forests

These are wetlands dominated by a species-rich forest

community growing on waterlogged peat. They are

botanically very rich and may be of great conservation

importance for amphibians and for some fish, such as the

walking catfish Prophagorus nieuhofi and possibly the highly

endangered Asian Bonytongue Scleropage fonnosus. In

terms of their avifauna, they are important in supporting

many arboreal members of the lowland forest community,

which are scarce or absent elsewhere due to the almost

complete destruction of terrestrial lowland forest. Phru Toh

Daeng or Phru Sirindhom, in Narathiwat Province, is the only

example of this habitat remaining in Thailand, although many

other areas, now dominated by the species-poor Melaleuca

woodland, may be degraded remnants of this type.

MANAGEMENT OF WETLANDS IN SOUTHERN

THAILAND

Wetlands are now known for many people in Southern

Thailand. So are their functions and values. Some

management practices have long been imposed on these

areas although not all wetlands are properly managed.

These practices may be categorized roughly, according to

responsible organizations, into the following five categories.

Conservation Imposed by the Government

Several wetlands, including many intertidal mangroves,

forested upstream wetlands, and some primary peat forests

have been designated by the government as conservation

areas. Among others, national parks, nonhunting areas, and

wildlife sanctuaries are most common. These practices

usually target objectives such as conserving wildlife, plant

species, and wetland ecology. Human activities have, to

some level, been controlled by the central government

through responsible departments such as the Royal Forestry

Department (RFD) and the Department of Fisheries (DOF).

However, ecotourism, which has been allowed and

promoted in many areas, may lead to more of the general

public entering these wetlands and thus posing a threat to
45

them. One good example of wetlands with this type of

management can be seen at Phru Toh Daeng in Narathiwas

province.

Management Imposed by Responsible

Government Agencies

Similar to the first category, this type of management,

usually invented by responsible authorities in line with the

government policy, aims at controlling the uses of the

wetlands where such agencies have been granted special

right to access and manage. Whereas the former strategy

usually aims to conserve the wetlands, these agencies,

Electricity Generating Authority of Thailand (EGAT) and

Royal Irrigation Department (RID), for example, could seek

to optimize their principal objectives while taking into

account conservation practices as secondary objectives

wherever possible. A constructed wetland complex at

Chewlam Dam in Suratthani is a good example of wetlands

with this type of management.

Management Introduced and Supported

by Local People

In line with the increasing awareness of the general public,

some wetlands have been conserved by various groups of

local people, many of whom rely to some extent on the

wetlands. Without support from the government, this type of

management may lead to resource-use conflicts because

some other groups sometimes want to make different uses

of the same wetlands. It is, however, believed to be most

cost effective and more in line with new legislation and

modem resource management paradigms—the integrated

and participatory approach, in particular.

Also, there are several projects funded by overseas

organizations also suggesting a similar approach. Whereas

the first strategy will likely remain essential, this

management practice is gaining more acceptance. A good

example of wetlands with this type of management is

Talenoi, Phattalung. There are several other examples in

Songkhla Lake Basin and some intertidal mangrove areas in

Suratthani and Pattani.

Owned by the Government but Accessible

Freely by the General Public

This type of management is common in most midstream

wetlands such as flood plain grass swamps and some peat

forests, most of which are of lower value compared to the

other natural wetlands. The government usually designates

these areas as public lands but has paid less attention to

control of their utilization. As a result, it is considered less

efficient and often leads to misuses or deterioration of these

wetlands or both. These practices also result in many

conflicts either among different users or between villagers

and the government. Several grass wetlands in Suratthani

and Nakhonsrithammarat are obvious examples of these

wetlands.

Experimental Management by Researchers

Several action research projects have been conducted by

many organizations including universities, international

organizations, and some government agencies. Important

research teams include, among others, Prince of Songkla

University, Kasetsart University, Wetlands International,

ASEAN, USAID-CRMP, DANCED, National Research

Council of Thailand, Department of Fisheries, and Royal

Department of Forestry. These research projects usually led

to a set of management guidelines, some of which were

adopted by responsible authorities. Some also led to actual

management practices based on scientific data and

reasoning. Local people also support many research

outputs.

Many organizations including responsible government

agencies, universities, nongovernment organizations, local

administrative organizations, and local people have played

different roles in wetlands management, sometimes leading

to resource-use conflicts. A more integrated approach to

management, which takes into account different people or

organizations or both, their expectations, responsibilities,

and underlying legislation have gained more acceptance

and would likely result in restructuring of the region’s

wetlands management scheme. The Songkhla Lake

management is one good example of the new paradigm.

Research into Wetlands in Southern Thailand

Numbers of research into intertidal wetlands in Southern

Thailand were and have been carried out using satellite

imagery, aerial photography, ground surveys, and laboratory

testing to examine various aspects of major wetlands in the

region. Among many, the extent of wetlands, the

ecosystems, waterfowls, and other biological species have

been known to some degree. Research into the only

remaining primary peat forest at Phru Toh Daeng has also

been carried out, and an integrated management plan for

the site was developed. Several aspects of Songkhla Lake,

the largest and most important lagoonal wetland complex,

were and have also been studied. Many management

proposals have been presented.

Apparently, past research into wetlands in this region had

paid more attention to estuarine wetlands, mangroves in

particular. More recent research has also extended to cover

rain-forested wetlands and some peat forests. Most of the

midstream wetlands, such as flooded grasslands, have been

left unexplored. So have many small upstream and isolated

wetlands.

Different research teams have collected various information

about wetlands flora and fauna. Research interests have,

however, focused upon the wetlands being habitats for

wildlife, various bird species in particular. Some research

into proper management has been carried out. Most

research in the past aimed at each particular wetland

although many are interrelated and could affect one another.

A few projects applied systems approach to wetland

research. Songkhla Lake management is among those who

based their investigation on the wetlands system.

Just recently, Prince of Songkla University, in cooperation

with Wetlands International, has set up a group of research

projects aiming at wetlands inventory and management in

Southern Thailand. Only a few issues have yet been handled

so far.
46

RECOMMENDATIONS

The authors suggest that wetlands in Southern Thailand,

based on recent investigation, should be categorized into

three groups.

Mountainous Upstream Riverine Wetlands

These wetlands should include upper perennial rivers and

hilly water storage reservoirs.

Lowland Midstream Riverine Wetlands

These wetlands should include lower perennial rivers,

freshwater lakes, ponds and associated marshes, freshwater

swamp woodlands, lowland water storage reservoirs, and

rice paddies.

Lowland Coastal and Marine Wetlands

These wetlands should include open sea coast, sandy

beaches and offshore islands, intertidal mudflats and

mangroves, and peat swamp forests.

The new scheme takes into consideration the system of

wetlands and thus should emphasize the interrelationships

among different wetlands, which call upon more systems

approach to management and further research.

Future research into wetlands should extend from lowland

intertidal mudflats and mangroves and a few major upstream

wetlands to peat forests, both intact and degenerated, and

lowland midstream wetlands, which also support various

forms of wildlife, natural conservation, and human activities.

Issues for investigation should extend from biology or

ecology or both to include human activities, problems and

their causes, and appropriate management practices that

would lead to sustainable wetlands in this region.

REFERENCES

Damsawasdi, R. 2000. Field investigation of wetlands in Southern

Thailand. Hatyai, Thailand: Prince of Songkla University. 52 p.

Scott, Derek A., ed. 1989. A directory of Asian wetlands. Gland,

Switzerland; Cambridge, U.K.: IUCN, The World Conservation

Union: 667–736.
47

THE DEVELOPMENT OF A DECISION SUPPORT SYSTEM FOR PRIORITIZING

FORESTED WETLAND RESTORATION AREAS IN THE LOWER YAZOO RIVER

BASIN, MISSISSIPPI

Angela A. Davis, Barbara A. Kleiss, Charles G. O’Hara,

and Jennifer S. Derby1

1

 Davis, GIS Specialist, U.S. Department of Interior, Geological Survey, National Wetlands Research Center, 700 Cajundome Boulevard,

Lafayette, LA 70506; Kleiss, Supervisory Hydrologist, and O’Hara, Hydrologist, U.S. Department of Interior, Geological Survey, 308 South

Airport Road, Pearl, MS 39208; and Derby, Ecologist, U.S. Environmental Protection Agency, 61 Forsyth Street, Atlanta, GA 30303.

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability of

wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.

Abstract—The Eco-Assessor, a GIS-based decision-support system, has been developed for the lower part of the Yazoo

River Basin, Mississippi, to help planners and managers determine the best locations for the restoration of wetlands based

on defined ecological and geographic criteria and probability of success. To assess the functional characteristics of the

potential restoration areas, the data layers are organized by hydrology, water quality, and habitat. The overall potential

restorability, or the predicted physical ability of a tract of land to sustain a functional wetland, is also considered. Because

an exact spatial representation of wetlands in the lower Yazoo River Basin does not exist, surrogate data layers are used to

predict locations that might be restored to a functional wetland. The Eco-Assessor analyzes the following data layers by

using a ranking system: geomorphology, soils, mature forest cover, farmed wetlands, flood frequency, topographic depressions, River Reach File Level 3 streams, wildlife management areas, conservation areas, primary roads, secondary roads,

permanent water, and landscape factors. Various categories of each data layer are assigned a rank. A higher rank indicates

that a particular geographic area has a higher probability of being restored to a functional wetland. Ranks for all the data

layers are summed to result in a cumulative rank which can then be used to determine the areas that, overall, are the most

likely to be successfully restored to a functional wetland. The ranking system method provides a means to analyze various

restoration scenarios. A restoration scenario can be defined in a way that may focus equally on all functions, focus on one

function, or focus on a particular geographic area.

INTRODUCTION

Forested wetlands, once the predominant land cover on the

Mississippi River Alluvial Plain (Creaseman and others

1992), provide habitat for wildlife, water-quality benefits,

flood storage, and many other ecological and environmental

benefits. Ongoing efforts of many Federal, State, and local

agencies and organizations to restore forested wetlands

have been successful. However, the lack of quantitative

methods for prioritizing the selection of wetland restoration

areas has meant that a less than optimum approach has

been taken in the evaluation, selection, and restoration of

forested wetlands.

In the past, selecting areas for wetland restoration was

conducted based largely on identifying landowners willing to

sell their land. This selection method, coupled with the lack

of a quantitative approach for selecting and prioritizing

potential restoration sites in past efforts, caused the process

of forested wetland restoration to overlook how the

restoration activity occurred on the landscape. Also, forested

wetland “restoration” was often undertaken with little regard

as to whether wetland functions were replaced. Until

recently, the evaluation of alternate forested wetland

restoration scenarios was a task that was impeded by the

general unavailability of input data, the cost of developing

digital data, the lack of sufficient tools for developing and

comparing alternate scenarios, and the difficultly of

integrating results into various types of independent

analysis. Recent improvements in data availability,

geographic information system (GIS) applications, computer

technology, and general software technology have made

possible the development and use of powerful decisionsupport systems (DSS) that integrate data, provide flexible

analysis methods, and allow the easy interchange of data

between various software analysis tools.

The DSS presented in this paper is the result of an

interagency agreement between the U.S. Department of the

Interior, Geological Survey, Water Resources Division, and

the U.S. Environmental Protection Agency. The purpose of

the agreement was to develop a DSS to facilitate the rapid

generation and consideration of many alternate forested

wetland restoration scenarios for the study unit located in

the Yazoo River backwater area.

STUDY AREA

The study area examined was the Yazoo backwater area of

the lower Yazoo River Basin, which included at least a

portion of the following six counties in Mississippi: Warren,

Issaquena, Sharkey, Yazoo, Humphreys, and Washington.

The Yazoo backwater area is in the southern portion of the

Yazoo River Basin, bounded by the Mississippi River levee

on the west and the valley wall on the east and south. The

southern tip of the Yazoo backwater area is just north of

Vicksburg, MS. The area extends north about 100 km to a

latitude near Belzoni, MS (fig. 1).
48

PROCEDURES

Ecological Rule Development

Wetland functions are often used as comparative evaluation

criteria for the ecological merit of wetlands.

Hydrogeomorphic Assessment (HGM) uses wetland

functions to evaluate existing wetlands at a site-specific level

(Brinson 1993). The Eco-Assessor DSS uses similar

principles to HGM, but spatial data at a landscape level is

grouped by wetland functions to evaluate whether or not a

sustainable wetland could exist at a location where there

currently is not one. In this study, common wetland functions

have been grouped into four categories: restorability,

hydrology, water quality, and habitat functions. For the

purposes of the initial model generation, the hydrology,

water quality, and habitat function categories were given

approximately the same weight in the overall analysis,

whereas restorability has a weight of approximately half that

of the other three. However, the model is adaptable, using

check boxes and pull-down menus, which allow the model to

be run with any preferred subset of functional categories

emphasized, and it also allows the input of new ranking

values for any of the functions. This kind of customization

can address particular resource needs or test hypotheses

about the impact of ranking decisions and weights. Also in

some cases, a particular function appears in more than one

functional category. In these cases, the function is

considered critical enough that the redundancy is justified.

Restorability—This functional category is defined as the

physical ability of a parcel of land to sustain a functional

wetland. The wetland restorability section of the EcoAssessor DSS provides for the evaluation of geomorphology,

soils, regeneration distance, and farmed wetlands spatial

data layers. A summary of the ranks assigned to functions

within this group is presented in table 1.

Geomorphology, as derived by Saucier (1994), consists of

abandoned channels, backswamps, and pointbar/valley

trains. Abandoned channels are considered the lowest land

formations in terms of elevation, become inundated most

frequently, and are given the highest rank. Backswamps are

Figure 1—The boundaries of the study area located in the Lower

Yazoo River Basin in Mississippi.

Table 1—Summary of the ecological rules used in the Eco-Assessor 

decision-support system for the wetland restorability function 

 Functional 

Wetland restoration

function Spatial data layer Data variables ranking 

Wetland 

restorability Environment of 

 deposition Abandoned channel 5 

 Backswamp 3 

 Pointbar 1 

Soils Hydric 10 

 Nonhydric 1 

Regeneration 

distance Within 60 m of 

 mature forest 5 

 Between 60 and 120 m 3 

 Greater than 120 m 1 

NCRS 

farmed wetlands Farmed wetland 5 

 Other 0 

NCRS = Natural Resource Conservation Service. 
49

slightly higher in elevation than abandoned channels but are

still low enough to be frequently inundated. Pointbar/valley

trains are slight ridges on the land’s surface, are the least

wet, and are given the lowest rank.

Hydric soils are defined by the U.S. Department of

Agriculture, Natural Resources Conservation Service

(NRCS), and modified for the Yazoo River Basin by using the

U.S. Army Corps of Engineers report “Delineation of

Wetlands of the Yazoo River Basin in Northwestern

Mississippi,” Misc. Paper EL–92–2 (Kirchner and others

1992). The presence of hydric soils is important to the

sustainability of wetlands.

Restoration areas that are near existing mature forest tend

to have much higher species diversity than areas that are far

from an existing stand of mature forest (Allen 1990). The

rapid natural regeneration of forest will occur out to a

distance of 60 m from existing forest (Allen 1997), and areas

within 60 m of existing forest are given the highest ranking in

this model.

The criterion for NRCS “farmed wetlands” is for areas

(excluding pothole, playa, and pocosin) that have a 50-

percent chance of being flooded or ponded for at least 15

consecutive days during the growing season (U.S.

Department of Agriculture 1996). Areas classified as NRCS

“farmed wetlands” indicate places on the landscape that are

inundated for a significant duration and are, therefore, very

likely to maintain sustainable wetlands and are given a high

rank.

Hydrology—The hydrology function is a representation of

the hydrologic regime of a given cell within the landscape.

Both the frequency and duration of flooding are considered.

The hydrology section includes and provides for the

evaluation of flood frequency and topographic sinks spatial

data layers. A summary of the ranks assigned to functions

within this group is presented in table 2.

Those areas indicated as flooded by a 0.5-year flood are the

most frequently inundated and, therefore, most likely to

sustain a wetland. The areas within the 2-year flood are not

inundated as often but are still viable sites for a wetland.

Those areas beyond the 2-year but within the 100-year flood

are the least likely to be inundated on a regular basis and

are the least likely areas for a sustainable wetland. The flood

frequency data were created by compositing the nominal

flood image data for the 0.5-, 2-, and 100-year nominal flood

images into a composite nominal flood frequency image

(fig. 2).

Topographic depressions indicate places on the landscape

where water is likely to pond because they are points of low

elevation surrounded by points of higher elevation. Once

water enters a depression, there is no outlet through which

the water is able to drain. The lack of an outlet causes the

Table 2—Summary of the ecological rules used in the Eco-Assessor 

decision-support system for the wetland hydrology function 

 Functional 

Wetland restoration

function Spatial data layer Data variables ranking 

Hydrology Flood frequency Within the 0.5-year flood 20 

 Within the 2-year flood 10 

 Beyond the 2-year flood 5 

Topographic 

depressions Topographic depressions 20 

 Other 1 

Figure 2—The spatial extent of the 0.5- and 2.0-year floods

in the Lower Yazoo River Basin, as determined from satellite

imagery
50

water to remain in the sink until evaporation or seepage or

both occur, resulting in floodwater storage and possible

interaction with ground-water resources.

Water quality—The water-quality function gives weight to

areas on the landscape that would filter, trap, or degrade

chemical components such as nitrogen and phosphorous

commonly found in the water. The water-quality section

includes and provides for the analysis of spatial data layers

such as stream buffers, flood frequency, and topographic

sinks. A summary of the ranks assigned to functions within

this group is presented in table 3.

Flood frequency is a factor in both the wetland hydrology

function as well as the wetland water-quality function. In the

water-quality function grouping, those areas indicated as

flooded by a 0.5-year flood are the most frequently

inundated. Therefore, the areas within the 0.5-year flood are

the most likely to sustain a wetland. The areas within the 2-

year flood are not inundated as often but are still viable sites

for a wetland. Those areas beyond the 2-year but within the

100-year flood are the least likely to be inundated on a

regular basis and are the least likely areas for a sustainable

wetland.

Stream buffers are assigned by stream level beginning with

a 10-m buffer because no less than a 10-m stream buffer is

a minimum necessary to filter nitrogen and phosphorous

(Dillaha and others 1989, Howard and Allen 1988). Stream

buffers have been shown to control the flow of nitrate,

phosphorous, sediment, and sediment-borne chemicals in

surface runoff and shallow ground water (Lowrance and

others 1997).

Topographic depressions allow water to pond in areas with

no outlet through which water can drain. If water remains

trapped in a topographic depression for extended periods of

time, sediments fall out and anaerobic processes begin. The

amount of sediment that will be deposited in depressional

areas is higher than in nondepressional areas because

longer hydroperiods allow for longer settling time (Hupp and

Morris 1990, Kleiss 1996). Both the trapping of the

sediments and the degradation of chemicals through

anaerobic processes improve overall water quality (Mitsch

and Gosselink 1993).

Habitat—The habitat function gives weight to areas of the

landscape in which wildlife may easily persist. The habitat

section considers proximity factors, such as distance to

wildlife management areas and conservation areas, distance

away from primary and secondary roads, proximity to

permanent water bodies, and landscape factors such as

forest block size and core area. A summary of the ranks

assigned to functions within this group is presented in

table 4.

Proximity Functions

The public lands are divided into two categories. The first

category contains the managed wildlife areas, National

Wildlife Refuge And State Wildlife Management Areas. The

second category contains general conservation lands,

Public Land Restoration, Delta National Forest, Farmer’s

Home Administration, and Wetland Reserve Program lands.

Expanding existing public lands, when managed

appropriately, greatly benefits wildlife by increasing the

interior space available for habitat. Also, any connections

that can be made between two patches of land add valuable

corridors for the movement of wildlife (Allen and Kennedy

Table 3—Summary of the ecological rules used in the Eco-Assessor 

decision-support system for the wetland water-quality function 

 Functional 

Wetland restoration

function Spatial data layer Data variables ranking 

Water quality Flood frequency Within the 0.5-year flood 15 

 Within the 2-year flood 10 

 Beyond the 2-year flood 5 

Stream buffers Stream level 1: within 90 m 15 

Stream level 2: within 80 m 15 

Stream level 3: within 70 m 15 

Stream level 4: within 60 m 15 

Stream level 5: within 50 m 15 

Stream level 6: within 40 m 15 

Stream level 7: within 30 m 15 

Stream level 8: within 20 m 15 

Stream level 9: within 10 m 15 

Stream level 0: within 5 m 15 

 Other 0 

Topographic 

depressions Topographic depressions 15 

 Other 0 
51

1989). The expansion of existing wildlife management areas

has an added benefit because the management of wildlife is

already the top priority in this area.

Distances away from primary and secondary roads were

adapted from a Louisiana Department of Natural Resources

study by Kinler (1994). The study ranked human

disturbances by distance and type of disturbance. For the

purposes of the Eco-Assessor, primary roads are

considered a constant disturbance and secondary roads are

considered only a frequent disturbance.

Being near permanent water bodies is beneficial to wildlife

because water is a requirement for basic living needs. In a

study conducted in the same general geographic area

(Wakeley and Marchi 1992), six species were chosen for a

habitat evaluation of the Upper Steele Bayou area in

Mississippi. The six species, which are common to

bottomland hardwood forest, include the barred owl (Strix

varia Barton), gray squirrel (Sciurus carolinensis Gmelin),

Carolina chickadee (Parus carolinensis Audubon), pileated

woodpecker (Dryocopus pileatus Linnaeus), wood duck (Aix

sponsa Linnaeus), and mink (Mustela vison Schreber)

(Wakely and Marchi 1992). Of these six species, the pileated

woodpecker has the most quantitatively specific habitat

requirements according to the U.S. Fish and Wildlife Service

Habitat Suitability Index Model (HSI). Minimum distance

requirements to and from permanent water bodies, as well

as minimum forest block size, are given in the HSI. The HSI

for the pileated woodpecker indicates that nesting habitat

generally is not observed greater than 150 m from water

bodies (Schroeder 1982). The habitat requirements for the

pileated woodpecker are often used as a representation of

the habitat requirements for other cavity nesting birds by

natural resource agencies (Renken and Wiggers 1993).

Landscape Factors

The landscape can be assessed using such landscape

factors as patch size, core area, and patch shape. A patch of

forested land < 1 ac does not provide sufficient habitat for

wildlife (Wakely and Marchi 1992); therefore, any patch < 1

ac is dissolved. The larger the patch size the more that

Table 4—Summary of the ecological rules used in the Eco-Assessor decisionsupport system for the wetland habitat function 

 Functional 

Wetland restoration

function Spatial data layer Data variables ranking 

Habitat Wildlife management areas Within 250 m of wildlife 

 management areas 10 

 Between 250 and 500 m 5 

 Between 500 and 1000 m 1 

 Beyond 1000 m 0 

Conservation areas Within 60 m of 

 conservation areas 5 

 Between 60 and 120 m 3 

 Between 120 and 500 m 1 

 Beyond 500 m 0 

Primary roads Within 50 m of 

 primary road 0 

 Between 50 and 500 m 1 

 Beyond 500 m 3 

Secondary roads Within 50 m of 

 secondary road 1 

 Between 50 and 500 m 2 

 Beyond 500 m 3 

Permanent water Within 150 m of 

 permanent water 5 

 Between 150 and 1000 m 1 

 Beyond 1,000 m 0 

Forest block size Between 1 and 10 acres 1 

 Between 10 and 320 acres 5 

 > 320 acres 10 

Core area ratio Ratio of core area to total 

 area of patch > 0.66 10 

 Between 0.33 and 0.66 5 

 < 0.33 1 
52

habitat is benefited. There are two categories of wildlife

species: generalists and specialists. Generalists can live in

patches of many shapes and sizes because their

populations are larger and they are highly mobile. It is the

specialists that require the greatest conservation efforts.

Specialists require large patches of forest with greater

interior area and less edge (Kinler 1994). It is important to

provide for the needs of the specialists by giving weight to

larger patches of land.

The ratio between core area and total patch area is used to

give more weight to patches of land that have a greater

portion of interior area. Core area is defined by the Fragstats

manual as “the area within a patch beyond some specified

edge distance or buffer width” (McGarigal 1995). Any land

that is in the interior of a patch more than 100 m from the

edge is considered core area. For a given patch of land, the

number of cells considered core area divided by the number

of cells in the entire patch, results in a ratio of core area to

patch shape. A long thin patch of land would receive a lower

ratio, whereas a long wide patch of land would receive a

higher ratio. A patch of land with a high ratio would provide

wildlife habitat with fewer edge effects and more interior

space. More interior space available in a given patch gives

rise to the number of interior species and species diversity

(Ohman and Eriksson 1998).

GIS Data Layers

The scope of this project did not include the collection of

new data to develop new data layers; therefore, the data

layers are the best information currently available from

agencies working within the State of Mississippi. In some

cases, data layers are numerically derived from existing

layers. All data layers were resampled to generate a grid of

25-m cell size.

Land use image data—In 1988, the U.S. Army Corps of

Engineers, Vicksburg District (USACE), collected satellite

image data for the purpose of generating a land use

classification data set. Land use in the study area was

based on these data and was provided by the USACE in the

Universal Transverse Mercator (UTM) projection, North

American Datum (NAD) 27. The satellite image data were

divided into the following land use categories: cotton,

soybeans, corn, rice, herb1, herb2, pasture (grass),

bottomland hardwood, swamp, river, lake, and pond. In

areas within the satellite images where the land use was

obscured by cloud cover or where the spectral response is

similar to that provided by sandbars, selected pixels are

classified as sandbar/clouds. For areas not classified, null

data values indicate the spatial extent of the study unit.

After generating land use data, the USACE adjusted the

land use. The adjusted land use data differ from the original

land use classifications where it is known the land has been

put into habitat or land use management programs such as

wildlife management areas, wetland reserve program lands,

national wildlife refuges, and conservation reserve program

lands.

Flood image data—The USACE collected satellite imagery

of flood scenes to accomplish several specific tasks. The

primary tasks included compiling areas inundated by floods

of a known stage to develop a stage-area relation and

spatially characterizing areas inundated by a flood of a given

frequency.

A relation between flood stage and inundated area is

developed in the form of a stage-area curve by selecting

images for flood scenes of various stages and determining

areas inundated. The stage-area curve is useful in

estimating flood extent. Dates and stages for flood scenes

used by the USACE in generating a stage-area relation for

the lower Yazoo River Basin are listed in table 5. By

comparing those dates and times with satellite image data

availability, it was possible to select images that

corresponded to specific flood events for areas in the vicinity

of specific gages within the study area. Image data for each

gaged area for flood events of specific frequency were

composited into a mosaic of images. The nominal flood

image data generated from composited images provided a

view of a simulated flood in which all areas are at the stage

for the specific flood considered. The 2-year nominal flood

image scene is shown in figure 2.

Hydric soils data—The soils data mapped and provided by

the USACE show the extent of hydric, nonhydric, and

riverbottom soils throughout the study area (fig. 3). The

presence of hydric soils in a location provides one of the

physical indicators that the location might support wetland

function. Hydric soils, as defined by the NRCS were for the

Yazoo River Basin in northwestern Mississippi (Kirchner and

others 1992).

Geomorphology data—The base source for hydrogeomorphology GIS data compiled at 1:250,000 scale was the

report “Geomorphology and Quaternary Geologic History of

the Lower Mississippi Valley” (Saucier 1994). This data layer

provides an indication of the fluvial environment that gave

rise to specific landforms and divides the landscape into

areas that are characterized as abandoned channels,

backswamps, and pointbar/valley trains.

Public lands data—The public lands data were obtained

from the U.S. Department of the Interior, Fish and Wildlife

Service (USFWS). The data contain six public land types,

including Public Land Restoration, Delta National Forest,

Farmer’s Home Administration, Wetland Reserve Program,

National Wildlife Refuge, and State Wildlife Management

Area lands.

Roads and transportation data—Data for primary and

secondary roads and railroads were obtained from the

Mississippi Automated Resources Information System

(MARIS). The sources for MARIS transportation data layers

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

digital line graph (DLG), and the Mississippi Department of

Transportation. The 1:100,000-scale primary roads data

include interstates, U.S. highways, and 1- and 2-digit State

highways; for example, Mississippi Highway 3 and Highway

25. The secondary roads layer includes 3-digit highways; for

example, Mississippi Highway 471 and the Natchez Trace

Parkway.

Water bodies data—Permanent water bodies were adapted

from MARIS permanent water data. This data layer was
53

modified by removing areas designated as catfish ponds, for

the specific needs of this effort.

Elevation contour data—Hypsographic contours were

made available through MARIS as compiled from 1:24,000-

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

base material.

Hydrologic network data—The EPA River Reach File Level

3 (RF3) was selected to represent the hydrologic network for

the area. Stream buffers were created from the RF3 streams

network coverage by using stream level. Stream level ranges

from 0 to 9. A level 1 stream flows to the ocean. A level 9

stream would be the size of a small creek or a ditch. A level

of 0 indicates that the actual stream level is unknown in the

RF3 dataset.

NRCS “farmed wetlands”—The NRCS created the “farmed

wetlands” data set such that the criteria for “farmed

wetlands” was for areas (not pothole, playa, or pocosin) that

have a 50-percent chance of being flooded or ponded for at

least 15 consecutive days during the growing season. Areas

classified as NRCS “farmed wetlands,” indicate places on

the landscape that are inundated for a significant duration

and are, therefore, very likely to maintain sustainable

wetlands. The NRCS “farmed wetlands” are defined for

regulatory purposes and are not indicative of farmed areas

that may be considered wetlands under alternate wetland

definitions.

Table 5—Stage measurements for seven stream gauging sites in the Yazoo backwater 

areaa

 

 Big 

 Steele Sunflower Big Little 

Steele Steele Bayou River at Big Sunflower Sunflower

Bayou Bayou at Little Sunflower River at River at 

Date of at at Upper Callao River at Holly Upper 

image Grace Onward Intake Landing Anguilla Bluff Intake 

03/12/73 93.3 85.5 77.2 97.1 94.1 89.5 82.2 

03/31/73 98.2 92.5 89.3 102.8 99.3 96.5 95.4 b

05/05/73 100.4 100.6 100.2 101.9 100.7 100.4 100.3 

01/30/74 98.3 92.9 90.6 101.2 98.1 96.0 93.4 

01/13/83 94.6 93.1b 91.9 100.8 98.1 95.5 93.1 

02/17/84 92.6 85.8 76.1 99.4 94.3 90.5 81.4 

03/05/87 91 84.9 79.5 98.7 94.7 90.0 82.4 

12/02/87 86.9 73.7 66.2 87.0 83.0 79.1 70.8 

03/10/89 89.7 89.7 b 89.7 90.1 89.0 91.5 90.0 

04/01/91 87.5 N/A 83.3 89.4 87.7 86.0 83.8 

04/30/91 98.1 93.9 90.4 103.0 98.5 95.4 91.7 

06/04/91 86.9 85.2 84.8 95.0 93.8 92.4 89.1 

02/01/93 86.3 83.4 83.0 84.7 83.8 83.6 83.2 

All measurements are in feet above mean sea level. 

N/A = not available. 

a Stage values are from U.S. Army Corps of Engineers published data except where indicated. 

b Dave Johnson, U.S. Army Corps of Engineers, written communication, November 2, 1998.

Figure 3—The spatial extent of hydric and nonhydric soil

in the study area.
54

Digital elevation data—High-resolution hypsographic data

were obtained from a collaborative effort with MARIS. Highresolution, drainage-reinforced digital elevation model (DEM)

data were developed by using the high-resolution

hypsographic data and the ArcInfo routine TOPOGRID,

which uses line information to create customized elevation

grids. The hypsography data layers were combined, and a

seamless elevation grid was created at a 10-m posting

interval (raster cell spacing) for the entire lower Yazoo River

Basin. A filled DEM layer and other hydrologic derivatives

were produced to allow the analysis of hydrology within the

study area. The filling of a DEM involves an automated

process wherein localized depressions (which in nature fill

and overflow) are digitally filled to provide a continuous

hydrologic surface.

The difference between the filled and the unfilled highresolution DEM was used to create a topographic

depressions data set (fig. 4). This layer provides an

analytical tool for assessing the size, distribution, and nature

of areas that can be considered as topographic depressions

and likely sites for water storage and functional wetlands

restoration.

Topographic base data—To provide a continuous

topographic base for the study area, USGS digital raster

graphics (DRG) of the 1:24,000-scale quadrangle maps

were collar-clipped (the white map collar or border was

clipped out), edge-matched, and placed into seamless

image catalogs for the study area. This topographic base

layer was used to provide quality assurance for all GIS data

layers used in the study.

Model Development

The Eco-Assessor program was written using Arc Macro

Language programming and runs using ArcInfo in a

Windows NT environment. Clickable menus are provided in

order to give the user the ability to turn on and off each data

layer as well as change the ranks assigned to each data

layer.

RESULTS AND DISCUSSION

Generation of “Functional Restoration” Maximum

Once the Eco-Assessor has analyzed each data layer, the

ranks for all data layers are summed. The summation results

in a functional restoration (FR) rank for each cell of eligible

land. The functional restoration rank is used as the indicator

of which areas on the landscape are the most suitable for

wetland restoration and would perform wetland functions

well.

Functional restoration maximum is the grid generated by the

Eco-Assessor that contains the total FR value for each

eligible cell in the study area. The FR maximum assumes

that every eligible cell within the study unit is selected for

reforestation. The total FR value is the sum of the assigned

rank for each data layer of a given cell. The resultant FR

maximum spatial data layer has cells that have cumulative

ranks that range from 15 to 140. The highest ranked areas

are those that would be the most suitable for wetland

restoration and would be most likely to perform wetland

functions. This is depicted in figure 5.

Scenario Generation and Evaluation

Reforesting all eligible areas within the study area is an

unrealistic goal; therefore, specific subsets of the study area

are recommended for reforestation. These subsets of the

study area or scenarios are selected by using the GIS and

establishing spatial criteria. The justification for the spatial

criteria may be based on a number of reasons. A scenario

may be based upon targeting a particular wetland function, a

certain geographic area, a certain feature on the landscape,

or any other set of criteria that can be spatially determined.

The use of the FR rank becomes particularly important when

considering scenarios. The FR rank allows for the

comparison of scenarios on an ecological basis. The total

FR rank for a given scenario provides an indicator of the

ecological benefits to be gained by reforesting the area

specified by the scenario. The total FR rank for a scenario is

divided by the total number of acres for that scenario. This

calculation results in a FR per-acre score for that scenario.

The FR per-acre score can then be used to compare the

ecological merits of various scenarios (fig. 6).

In the hydrology scenario, the eligible areas inundated in the

USACE nominal 2-year flood scene are selected for

reforestation (fig. 7).

Figure 4—The spatial extent of topographic depressions

in the study area.
55

SUMMARY

The Eco-Assessor DDS provides a valuable tool to prioritize

the restoration of forested wetlands in the lower Yazoo River

basin. The data compiled, and the tools that are included in

the DSS, facilitate the rapid generation and consideration of

alternate restoration scenarios. The DSS can be used to

help develop reforestation plans that place wetland forest

communities in locations in the landscape where each

wetland has a high probability of developing into a functional

wetland system. Reforestation efforts can be targeted to

areas that would provide the highest ecological benefit for a

given economic investment. The DSS also provides a

method for systematically altering the buffer distances and

ranks assigned to each data layer through the use of the

interactive menus, which make up the Eco-Assessor

framework. The ability to change the ecological rules and

rank values allows the user to obtain the most appropriate

ranking for a given wetland system.

ACKNOWLEDGMENTS

The authors would like to acknowledge the contributions of

the following people to this project: Roland Viger, Physical

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

Denver, CO; Al Rea, Hydrologist, U.S. Department of the

Interior, Geological Survey, Oklahoma City, OK; Sridhar

Katragadda, GIS Specialist, U.S. Department of the Interior,

Figure 5—Functional Restoration Maximum—the data layer

generated by summing the assigned ranks for all data layers for

each 25-m cell in the study area. Darker areas represent areas that

have a higher probability of being restored back to a functional

wetland.

Figure 6—Various restoration scenarios can be compared by

comparing their sum of functional restoration values (A) and by

dividing the functional restoration value by the area to get a

functional restoration value per land area estimate (B).

Figure 7—An example of a restoration scenario derived from the

Functional Restoration Maximum layer in which all cells within the 2-

year flood are selected.
56

Geological Survey, Pearl, MS; and Leonard Shabman and

Laura Zepp, Virginia Polytechnic Institute and State

University, Blacksburg, VA. Reviews were provided by: Wade

Bryant, Regional Biologist, U.S. Department of the Interior,

Geological Survey, Norcross, GA; Craig Harvey,

Geographer-GIS Specialist, Carol Moss, Editor, and Larry

Slack, Supervisory Hydrologist, U.S. Department of the

Interior, Geological Survey, Pearl, MS; Greg Easson,

Assistant Professor of Geology and Geological Engineering

and Nolan Augenbaugh, Professor of Geology and

Geological Engineering, University of Mississippi, Oxford,

MS; and Tammy Charron, Editor, Johnson Controls World

Services, National Welands Research Center. Lance Cooper

and Michael Wade provided graphics assistance. A portion

of this project was used in partial fulfillment of requirements

for a Master’s degree at the University of Mississippi, Oxford,

MS.

REFERENCES

Allen, J.A. 1990. Establishment of bottomland oak plantations on the

Yazoo National Wildlife Refuge complex. Southern Journal of

Applied Forestry. 14(4): 206–210.

Allen, J.A. 1997. Reforestation of bottomland hardwoods and the

issue of woody species diversity. Restoration Ecology. 5(2):

125–134.

Allen, J.A.; Kennedy, H.E., Jr. 1989. Bottomland hardwood

reforestation in the Lower Mississippi Valley. Slidell, LA.

Stoneville, MS: U.S. Department of the Interior, U.S. Fish and

Wildlife Service, National Wetlands Research Center; U.S.

Department of Agriculture, Forest Service, Southern Forest

Experiment Station. 28 p.

Brinson, M.M. 1993. A hydrogeomorphic classification for wetlands

WRP–DE–4. Vicksburg, MS: U.S. Army Corp of Engineers,

Waterways Experiment Station. 79 p.

Creaseman, L.; Craig, N.J.; Swan M. 1992. The forested wetlands of

the Mississippi River: an ecosystem in crisis. Baton Rouge, LA:

The Nature Conservancy of Louisiana: 4–5.

Dillaha, T.A.; Reneau, R.B.; Mostaghimi, S.; Lee, D. 1989. Vegetative

filter strips for agricultural nonpoint source pollution control.

Transactions of the American Society of Agricultural

Engineering. 32(2): 513–519.

Kinler, Q. 1994. Habitat assessment model for fresh swamp and

bottomland hardwoods within the Louisiana coastal zone. Baton

Rouge, LA: Louisiana Department of Natural Resources. 10 p.

+3 appendices.

Howard, R.J.; Allen, J.A. 1988. Streamside habitats in southern

forested wetlands: their role and implications for management.

In: Hook, D.D.; Lea, R., eds. Proceedings of the symposium: the

forested wetlands of the Southern United States; 1988 July 12–

14; Orlando, FL. Gen. Tech. Rep. SE–50. Asheville, NC: U.S.

Department of Agriculture, Forest Service, Southeastern Forest

Experiment Station: 97–106.

Hupp, C.R.; Morris, E.E. 1990. A dendrogeomorphic approach to

measurement of sedimentation on a forested wetland, Black

Swamp, Arkansas. Wetlands. 10(1): 107–124.

Kirchner, W.N.; Kleiss, B.A.; Clairain E.J., Jr. [and others]. 1992.

Delineation of wetlands of the Yazoo River Basin in northwestern

Mississippi EL–92–2. Vicksburg, MS: U.S. Army Corps of

Engineers, Waterways Experiment Station. 25 p. + 3 appendices.

Kleiss, B.A. 1996. Sediment retention in a bottomland hardwood

wetland in eastern Arkansas. Wetlands. 16(3): 321–333.

Lowrance, R.; Altier, L.S.; Newbold, J.D. [and others]. 1997. Water

quality functions of riparian forest buffers in Chesapeake Bay

watersheds. Environmental Management. 21(5): 687–712.

McGarigal, K. 1995. Fragstats: spatial pattern analysis program for

quantifying landscape structure. Gen. Tech. Rep. PNW–351.

Portland, OR: U.S. Department of Agriculture, Forest Service,

Pacific Northwest Station. 122 p.

Mitsch, W.J.; Gosselink, J.G. 1993. Wetlands. New York: Van

Nostrand Reinhold. 722 p.

Ohman, K.; Eriksson, L.O. 1998. The core area concept in forming

contiguous areas for long-term forest planning. Canadian Journal

of Forest Research. 28(7): 1032–1039.

Renken, R.B.; Wiggers, E.P. 1993. Habitat characteristics related to

pileated woodpecker densities in Missouri. The Wilson Bulletin.

105 (1): 77–83.

Saucier, R. 1994. Geomorphology and quaternary geologic history

of the Lower Mississippi Valley. Vicksburg, MS: U.S. Army Corp

of Engineers, Waterways Experiment Station. 364 p.

Schroeder, R.L. 1982. Habitat suitability index models: pileated

woodpecker. FWS/OBS–82/10.39. Fort Collins, CO: U.S.

Department of the Interior, Fish and Wildlife Service. 15 p.

U.S. Department of Agriculture, Natural Resource Conservation

Service. 1996. National food security act manual. 3rd ed. U.S.

Department of Agriculture.

Wakeley, J.S.; Marchi, R.A. 1992. An evaluation of impacts to

terrestrial habitats resulting from flood control projects in the

Steele Bayou Basin, Mississippi. Vicksburg, MS: U.S. Army

Corps of Engineers, Waterways Experiment Station. 26 p.
57

INTRODUCTION

Last year, the Fish and Wildlife Service reported that the rate

of U.S. wetland loss has slowed to a rate 60 percent below

that experienced in the 1970s and 1980s. Jamie Clark,

Director of the Fish and Wildlife Service, explained that the

study shows that our Nation’s efforts to restore and protect

wetlands are making a difference. At a time when many

Federal regulatory programs are criticized as too expansive,

these protective efforts are increasingly occurring at the

State level. Many States maintain wetland protection

schemes, and their legislatures continue to consider

methods of increasing control over State wetlands. Combine

this factor with express invitations from the U.S. Congress

and the Executive Branch, and the signs indicate that the

momentum behind the State initiatives will likely continue.

While States have initiated their own wetland protection

schemes for decades, Congress formally invited States to

join the regulatory game under the Clean Water Act (CWA) in

1977. The CWA Amendments provided two ways for States

to increase responsibility by assuming some administration

of Federal regulatory programs: State programmatic general

permits and State assumption. State programmatic general

permits (SPGP) allow a State to become the sole permit

issuer under an existing State-permitting program for

projects that have similar characteristics and will have low

environmental impacts. But, Federal control is maintained for

permitting other projects. SPGP have gained popularity, with

27 States holding permits of this type. In contrast, State

assumption has been less popular, with only two States

adopting such a program. In most instances, State

assumption grants more permitting authority to States:

States may issue individual permits for projects that do not

have to meet the general permit requirements of similar

nature and low environmental impacts. But, State

assumption also places a heavier burden on the State with a

stricter application and approval process, a greater funding

obligation, and a larger regulatory responsibility.

This paper reviews the two regulatory options available to

States under the CWA, focusing on the advantages and

disadvantages inherent in each. It explores the reasons why

States may choose to adopt Federal regulatory

responsibilities and presents a guide for States considering

one or both of these regulatory options.

EVOLUTION OF WETLANDS REGULATION

In the 1700s, 221 million ac of wetlands existed in the United

States. A 1995 inventory showed < 101 million ac remaining.

Historically, a wetland was considered a nuisance, believed

to inhibit navigation and provide habitat for little more than

mosquitoes. Thus, the Federal government encouraged

draining and filling of wetlands throughout the 1800s. Under

the Swamp Lands Acts, the Federal government granted 15

Western States almost 65 million ac for “swamp reclamation,”

making drainage and filling wetlands a national policy.

In the late 1960s, the Federal government took greater

notice of the benefits of wetlands. It began regulating the

filling and dredging of wetlands under the authority of the

Rivers and Harbors Act of 1899, which prohibited excavation

from or fill to any navigable water of the United States

without a recommendation by the Chief of Engineers and

authorization from the Secretary of the Army. The Corps also

began to consider the ecological benefits of wetlands and to

make more protective decisions regarding permits.

In 1972, Congress enacted the Federal Water Pollution

Control Act (Clean Water Act, CWA) to eliminate the

discharge of pollutants in the waters of the United States,

including wetlands, by 1985. Pursuant to this goal, the CWA

prohibits all persons from discharging pollutants into waters

of the United States unless they have obtained and are

operating within the strictures of certain permits. If the

pollutants involve dredged or fill material, the permits are

issued by the Corps with the EPA maintaining a supervisory

role.

STATES ASSUMING RESPONSIBILITY OVER WETLANDS: STATE ASSUMPTION

AS A REGULATORY OPTION FOR PROTECTION OF WETLANDS

Kristen M. Fletcher1

1 Director, Mississippi-Alabama Sea Grant Legal Program, 518 Law Center, University, MS 38677.

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability of

wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.

Abstract—While States have initiated their own wetland protection schemes for decades, Congress formally invited States

to join the regulatory game under the Clean Water Act (CWA) in 1977. The CWA Amendments provided two ways for States

to increase responsibility by assuming some administration of Federal regulatory programs: State programmatic general

permits and State assumption. States are also active in conservation programs such as preserving and managing wetlands.

State programmatic general permits (SPGP) allow a State to become the sole permit issuer under an existing Statepermitting program for projects that have similar characteristics and will have low environmental impacts. SPGP have gained

popularity whereas, in contrast, State assumption has been less popular with only two States adopting such a program. In

most instances, State assumption grants more permitting authority to States but also places a heavier burden on the State

with a stricter application and approval process, a greater funding obligation, and a larger regulatory responsibility.
58

The CWA Section 404 is the primary Federal regulatory

program providing protection for the Nation’s remaining

wetlands. The EPA and the Corps jointly administer the

program. The Corps’ responsibilities include day-to-day

program administration, individual permit decisions,

jurisdictional determinations, development of policy and

guidance, and enforcement. The EPA’s responsibilities

include development and interpretation of guidelines for the

environmental criteria used in evaluating permit applications,

identifying certain exempt activities, reviewing and

commenting on individual permit applications, exercising the

authority to veto Corps’ permit decisions, and overseeing

administrative responsibilities of the State assumption

program.

The section 404 process includes a public notice and

comment period. After receiving public comments, the Corps

evaluates the application to decide if it contributes to

conservation, economics, aesthetics, fish and wildlife values,

flood protection, general public welfare, historic values,

recreation, land use, water supply, water quality, and

navigation. A public hearing may be held, or the Corps may

grant or deny the application outright. An application for an

individual permit must meet the requirements of a

comprehensive inquiry through the Corps’ public interest

review and its analysis of compliance with the EPA

guidelines. The EPA guidelines seem to require a rigorous

examination of the availability of practicable alternatives, and

prohibit the authorization of any project that would result in

significant adverse impacts. In reality, the Corps denies < 10

percent of individual permit applications.

The CWA also authorizes general permits on a State,

regional, or national basis. Rather than applying for an

individual permit, a person may qualify to discharge dredged

or fill material into waters of the United States under one of

these permits issued for projects with small impacts. General

permits decrease the administrative burden for the Federal

and State regulatory agencies but have come under

increasing criticism because of the potential to greatly

contribute to overall loss of wetlands, little by little.

Some argue that the CWA was enacted precisely because

the individual States lacked the political will to clean up their

waterways and protect key resources. But, 5 years after its

enactment, Congress amended the CWA authorizing

substantial State regulatory participation. The amendments

authorized States to take over the section 404 permitting

program from the Corps and also provided for general

permits to relieve pressure created by expanded Federal

jurisdiction and, in part, as an acknowledgment of a practice

that the Corps already was performing.

One of the reasons Congress amended the CWA was to

address concerns that the section 404 program was too

overwhelming for the Army Corps of Engineers to manage

and that funds were insufficient. Reduction of Corps

workload was a primary reason for providing delegation to

the States. Even though parts of the legislative history

confirm Congress’ intent to delegate a greater responsibility

to the States, it also indicates concern for the performance

of State programs. Advocates argue that States occupy the

best position to take the lead in wetlands protection because

States tend to be more responsive than Federal agencies to

local needs but are still removed from the influences of local

politics, providing better protection than local governments.

In addition to the congressional invitation of 1977, the

Executive Branch has encouraged reducing Federal

involvement and replacing Federal wetlands regulators with

States. Specifically, the Bush Administration preferred a

minimum level of Federal involvement, citing Federal

regulatory programs as burdensome. The Clinton

Administration called for greater State action in its Wetlands

Policy of 1993. Two of the five principles for the Federal

Wetlands Policy directly related to States increasing their

responsibilities: avoidance of duplication between regulatory

agencies and expansion of partnerships with State and local

governments.

States were given alternative opportunities to respond to

pressures on wetlands. Since the passage of the CWA,

States were encouraged to establish their own conservation

and permitting programs and to work as partners with the

Corps in order to manage wetland areas. Some State

programs predated the Federal protections of wetlands, and

many took action in favor of their coastal wetlands prior to

extending protection to upland wetlands. Other States joined

programs sponsored by the Federal government that

combine Federal, State, local, and private efforts at restoring

and preserving wetlands.

States may also protect against the filling of wetlands

through the CWA Section 401 water-quality certification and

under the consistency determinations of the Coastal Zone

Management Act. Under the CWA Section 401, a State may

veto or condition a Federal licensed or permitted activity that

may degrade water quality or aquatic habitat, including

wetlands. It requires that an applicant for a Federal permit for

activities that may result in a discharge into navigable waters

must receive a certification from the State to insure

compliance with State water-quality requirements. Section

401 gives State water-quality control over a wide range of

activities for which they otherwise might lack such authority,

including wetlands preservation, protection of wildlife habitat,

and protection of aesthetic and recreational values of

waterways. It also allows States to limit impacts on wetlands

without running its own regulatory program or operating

under an SPGP or assuming section 404 authority from the

Corps.

A State may also use the consistency requirement under the

Coastal Zone Management Act to limit Federal permitted

activities, which affect wetlands in a coastal zone. If a State

has an approved coastal zone management program, then

an applicant for a Federal permit whose activity may affect

any land or water use or natural resource in the coastal

zone, must obtain certification from the relevant State

coastal resources agency that the permitted activity

complies with the State program and will be conducted in a

manner consistent with the program. A State may

specifically designate wetlands as regulated areas if they fall

within the State’s coastal zone. For fill activities in these

wetlands, the State may deny certification if it finds the

activities inconsistent with its program. Generally, a project

cannot continue without such certification. But State wetland
59

protection schemes have powerful 404 alternatives. As

regulatory tools, SPGP and State assumption must be

analyzed and adapted to further wetlands protection, in

addition to trying to simplify the permitting process.

STATE PROGRAMMATIC GENERAL PERMITS

The State programmatic general permit is one type of

general permit issued by the Corps. Today, the Corps uses

general permits to authorize 80 percent of the regulated

activities. The Secretary of the Army issues general permits

on a programmatic, regional, or nationwide basis. To qualify

for a general permit, the project must meet the following

requirements: (1) the activities authorized under the general

permit must be “similar in nature;” (2) the activities may

cause only minimal adverse environmental effects when

performed separately; (3) the activities may have only

minimal cumulative adverse effect on the environment; (4)

the permit must be based on EPA guidelines; (5) the permit

must be limited to a 5-year life span; (6) the Secretary can

revoke or modify the permit if the authorized activities have

an adverse impact on the environment, or such activities are

more appropriately authorized by individual permits.

The general programmatic permit furthers the idea that State

and Federal regulatory programs should complement rather

than duplicate one another. Corps regulations define

programmatic permits as “a type of general permit founded

on an existing State, local, or other Federal agency program

and designed to avoid duplication with that program.” Under

the SPGP, the Corps in effect delegates to a State the

primary responsibility under section 404 for permit review of

the activities meeting the requirements of the SPGP. As a

general permit, the SPGP must comply with the

congressionally mandated requirements set forth in section

404(e).

Before authorizing an SPGP, the Corps must analyze the

State program upon which it is based. Often, because of the

limited scope of some State programs, an SPGP will not

necessarily cover the entire State in question. Instead, the

Corps refers to any general permit program based on a

State program to assure the protection of wetlands as an

SPGP.

The programs tend to follow two models in practice: those

using fixed criteria and those using a process of consultation

and review. The SPGP using fixed criteria called the “New

England model,” requires the State to place potential section

404 permit applications in one of three categories: the

nonreporting, screening, or individual permit categories. If

the State places a permit application within the nonreporting

category, the applicant does not have to inform the Corps of

the activities. Instead, the applicant is only responsible for

meeting State requirements: a State permit, State waterquality certification, and a consistency concurrence, if the

project is in the coastal zone. An activity with higher impacts

may fall in the screening category, which requires an

interagency screening, or the individual permit category,

which requires an applicant to go through the Corps’

individual permit application process.

The second type of SPGP uses a consultation and review

structure. When a State agency receives an application for a

State permit, it conducts a site visit of the area and produces

a field report or evaluation. The State then produces public

notices of permit applications, putting the Corps and other

Federal resource agencies on notice. One of these agencies

or the State agency may request that the Corps require the

applicant to seek an individual permit. If not, the State

agency can issue the permit. This consultation and review

process is often criticized for increasing the workload and

delays but potentially provides a better review of the project

site and more accurate prediction of impacts.

States with SPGPs cite greater control over wetlands as a

reason to take on this authority. Through an SPGP, the State

can control the permitting of those actions with minimal

impacts, does not have to rely on the Corps, and avoids

duplication for these permit applications. Finally, the SPGP

gives States an alternative avenue to control the fate of their

wetlands other than by assuming the 404 permitting

process.

The SPGP is a win-win for the Corps as well. By giving the

State the authority to review and issue or deny these

permits, the Corps can reduce duplication between State

and Federal regulatory programs and reduce the Corps’

regulatory workload without compromising, at least from the

Corps’ perspective, the overall effectiveness of section 404

and section 10 permit review in protecting wetlands. In

addition, most SPGPs specifically exclude activities affecting

sensitive areas such as endangered species habitat or

historic properties and provide for kick-out provisions if

necessary.

STATE ASSUMPTION

The second option for States is State assumption of the

Corps’ permitting authority. Many States perceive that

SPGPs provide adequate control of State wetlands without

assuming the responsibility offered under State assumption

whereas others view SPGPs as merely a stepping stone to

assumption, which can offer a State more regulatory

authority than a general permit.

Unlike SPGPs, the CWA and accompanying regulations

specify the requirements for assuming section 404 authority.

The EPA must approve a State’s application to assume the

404 permitting program. The statute requires the Governor

of the applicant State to submit a description of the program

to the EPA, along with a statement from the State attorney

general that the laws of the State “provide adequate

authority to carry out the described program.” A State must

submit to the EPA Regional Administrator the following six

items: (1) a letter from the Governor of the State requesting

approval of the State program; (2) a complete program

description; (3) an attorney general’s statement confirming

that the laws of the State provide adequate authority to carry

out the described program; (4) a memorandum of agreement

with the EPA Regional Administrator; (5) a memorandum of

agreement with the Secretary of the Army; and (6) copies of

applicable State statutes and regulations, including those

governing applicable State administrative procedures.

The attorney general’s statement must also include

certification that each agency responsible for administering

the State program has full authority to administer the
60

program within its category of jurisdiction. In addition, the

State as a whole must have full authority to administer a

complete State program. Finally, the statement should

include a legal analysis of the likelihood of a constitutional

taking as a result of the successful implementation of the

State’s program.

In order to assume, the State will enter memorandums of

agreement (MOA) with both the EPA and the Corps. The

MOA with the EPA must set out State and Federal

responsibilities for program administration and enforcement

including provisions specifying classes and categories of

permit applications for which EPA will waive Federal review

authority and provisions addressing EPA and State roles and

coordination with respect to compliance monitoring and

enforcement activities. The MOA with the Secretary of the

Army must include a description of the waters of the United

States within the State over which the secretary retains

jurisdiction and an identification of all general permits issued

by the secretary, the terms and conditions of which the State

intends to administer and enforce upon receiving approval of

its program, and a plan for transferring responsibility for

these general permits to the State.

The program description must include various essential

elements in order to be approved. First, the description must

explain the State’s permitting, administrative, judicial review,

and other applicable procedures. In addition, it must include

a description of the funding and manpower available for

program administration, a description of how the State will

coordinate its enforcement strategy with the Corps and EPA

for nonassumable waters or projects, a comparison of State

and Federal definitions of wetlands, and the extent of the

State’s jurisdiction, scope of activities regulated, anticipated

coordination, and the scope of permit exemptions, if any. The

EPA distributes the State’s program submission to the

Corps, Fish and Wildlife Service, and the National Marine

Fisheries Service. EPA has up to 120 days to approve or

disapprove the State’s program. Once the EPA approves the

State application, the Corps transfers to the State those

permit applications for projects in the State’s jurisdiction.

The EPA retains oversight authority and receives copies of

all permit applications. The State must notify the EPA of any

action that it takes with respect to such applications. The

EPA Administrator provides copies of the application to the

Corps, the Department of Interior, and the Fish and Wildlife

Service and must notify the State within 30 days if the

administrator intends to comment on the State’s handling of

the application. The State must then await comment before it

may issue the permit. If the EPA objects to the application,

the State may not issue the proposed permit but may

request a hearing before the EPA or alter the permit to

accommodate the EPA objections. If the State does not

request a hearing, the EPA transfers authority to issue the

permit to the Corps. Once in the Corps’ hands, jurisdiction

remains there.

Finally, the statute requires that EPA review any revisions to

the State wetlands program, determine whether such

revisions are substantial or not substantial, and approve or

disapprove the revisions. The EPA also maintains the

authority to withdraw approval of the program. If the

administration of the State program does not meet EPA

guidelines, the EPA may take corrective action and may,

within a reasonable time, withdraw approval of the program

and redirect authority to the Corps.

MICHIGAN AND NEW JERSEY ASSUMPTION

PROGRAMS

Two States, Michigan in 1984 and New Jersey in 1994, have

assumed permitting authority with mixed results. An analysis

of these two programs provides a look at permitting under

State assumption.

Michigan began wetland permitting even prior to

congressional authorization for State assumption of section

404 authority. In 1955, the Michigan State legislature passed

the Great Lakes Submerged Lands Act, authorizing the

Michigan Department of Natural Resources (MDNR) to

regulate dredge, fill, and construction activities in the State’s

coastal zone. In 1972, the Michigan legislature acted again

by passing the Inland Lakes and Streams Act authorizing

the State to regulate activities occurring up to the ordinary

high-water mark on Michigan’s inland wetlands. Four years

later, Michigan developed a MOA with the Corps for a Joint

Public Notice System to expedite the issuance of wetland

permits. When congressional authorization for assumption

followed the next year, Michigan was well placed to assume

the wetlands permitting program. The State then passed the

Goemaere-Anderson Act of 1980, forming the framework for

assumption and expanding the State’s wetland permitting

requirements to include those wetlands, which were not

subject to section 404 jurisdiction.

The MOA between the EPA and Michigan named the MDNR

as administrator of the Michigan wetlands program. The

program’s procedures are similar to those under the CWA.

The Michigan DNR has 14 districts in 3 regions to handle

State permitting, mitigation matters, wetland delineation, and

enforcement. Upon submission of a completed application,

the MDNR has 90 days to issue a determination on the

proposed project; if the MDNR fails to issue a determination

within that time frame, the proposed project is considered

approved.

Prior to assumption, New Jersey’s wetlands scheme was

comprised of four acts. First, the Hackensack Meadowlands

Development Commission of 1968 set up a permitting

scheme for activities within district boundaries. The Wetlands

Protection Act of 1970 regulated activities within the coastal

and estuarine wetlands and required applicants to obtain

permits from the New Jersey Department of Environmental

Protection (NJDEP). In 1979, the New Jersey legislature

enacted the Pinelands Protection Act to regulate wetlands

and other areas within the Pinelands National Reserve.

Finally, in 1987, the legislature passed the Freshwater

Wetlands Protection Act developed specifically with

assumption in mind. The act created the State’s wetland

regulatory program and allowed the State to create buffer

areas adjacent to designated wetlands that are subject to

active regulation.

New Jersey assumed 404 authority in the spring of 1994.

The NJDEP was named as the State authority over the

program. The MOA with the EPA requires the State to submit
61

monthly and yearly reports to the EPA for review. The

NJDEP is not subject to strict 90-day time constraints like

the Michigan DNR. The program requires the NJDEP to act

“in a timely manner”; in practice, permit processing often

takes up to 4 months.

FEDERAL PERMITTING VERSUS ASSUMPTION

Improving the efficiency of permitting is a high priority for

those States considering assumption. The Corps is often

criticized for slow responses on permits. States can impose

a strict time limit on their permitting agencies, as Michigan

has done by requiring turn around by the MDNR in 90 days.

States also can provide more manpower than the Corps,

evidenced by Michigan’s creation of 13 field offices

throughout the State, as compared with 4 Corps offices. With

more field offices, decisionmakers can be more readily

available to applicants and can be closer to the wetlands

actually under their jurisdiction.

An element of improving efficiency of the 404 program is to

reduce the ever-present regulatory duplication. Under a

State-assumed program, paperwork for the applicant is

reduced, and the State agency becomes directly responsible

for the application. By reducing duplication, the State can

also increase predictability of the application process. A

State can also consolidate several different wetland statutes

to reduce the burden on the regulated public and streamline

the process.

Finally, advocates argue that wetlands will receive better

protection under an assumed program for three reasons.

First, the State is in a better position to address regional and

local concerns about the conservation and use of wetlands

resources. Assumption supporters claim that State agency

regulators are more familiar with the treatment and use of

the regulated lands. Also, because the State assumes

permitting authority over a smaller square acreage of

wetlands than the Corps had been responsible for, a State

can potentially provide closer examination of cumulative

impacts. Advocates also claim that a State regulator will be

more aggressive than its Corps predecessor, and a State

can avoid the inconsistency problems that often plague

Corps regulators.

Once State policymakers determine that State assumption

and SPGPs are a better alternative for wetlands protection in

their State, they must determine which alternative to employ

in their State. An analysis of the advantages and

disadvantages to each alternative is essential to this

determination.

Control Over Wetlands Decisions

States that have taken over some aspect of the Federal

regulatory program, and those considering such action

indicate that increased control over permitting decisions

regarding State wetlands is a driving force in seeking more

regulatory authority. But “control” over permitting decisions

hinges on several factors: jurisdiction, Federal oversight, and

flexibility of the program.

Jurisdiction

Jurisdiction of an SPGP varies according to the State

program upon which it is based. An SPGP generally covers

wetlands, waters, and activities within the Corps’ Section

404 and Section 10 jurisdiction that are regulated by a State

wetlands program. Therefore, a State can design its SPGP

to best match its regulatory program but can also amend this

underlying program to alter its jurisdiction. SPGPs do not

include those activities and areas that are considered to be

of national or international concern and many exclude

specific activities that have a potentially higher impact, such

as new or expanded marinas, projects requiring an

environmental impact statement, and wetlands fills over

specific acreage. A comparison of the underlying State

programs reveals the degree to which SPGPs can differ.

For instance, the Maryland SPGP covers only section 404

activities affecting < 5 ac of nontidal wetlands based upon

the jurisdiction of the Maryland Nontidal Wetland Protection

Act. The North Carolina SPGP applies to all section 404 and

section 10 activities that receive prior approval from the

State based on the State’s Coastal Area Management Act

permit, a State dredge and fill permit, or a section 401 State

water-quality certification when there is a discharge into U.S.

waters. But the North Carolina SPGP only applies to its 20

coastal counties. Finally, the Massachusetts PGP applies to

section 404 and section 10 activities that receive prior State

approval under the Wetland Protection Act Final Order of

Conditions, a Public Waterfront Act waterways license or

permit, a section 401 State water-quality certification, or a

State coastal zone consistency determination. Also,

jurisdiction under the PGP is expressly limited to wetland fills

less than or equal to 1 ac in size. Finally, these programs

expressly exclude activities affecting navigation, national

wildlife refuges, forests, parks, components of the National

Wild and Scenic River System, and threatened or

endangered species and their critical habitats.

The MOA between Michigan and the EPA divided jurisdiction

as follows: Michigan acquired responsibility for all activities,

which require dredging, placement of fill, or construction in

inland waterways; the Corps (Detroit District) maintained

jurisdiction over activities, which require dredging,

placement of fill, or construction in the Great Lakes coastal

areas, connecting waters, navigable waterways and those

wetlands adjacent to navigable waterways, and under other

specific circumstances. Similarly, New Jersey gained

permitting authority over freshwater wetlands except that a

1,000-ft boundary from the mean high water line of

navigable waterways was established as the jurisdictional

boundary between “adjacent” and “nonadjacent” wetlands.

As the above examples show, the two States that assume

section 404 permitting authority have jurisdiction over a

greater acreage of State wetlands. This does not necessarily

benefit the States, however. Jurisdiction under State

assumption can be changed if a State fails to meet EPA

guidelines. Thus, even though a State begins its 404

authority responsible for a large number of wetlands, the

Corps can reassume permitting authority over certain areas.

Whereas SPGPs can also be subject to change, especially

because they must be reapproved every 5 years, they offer

more flexibility. A State can easily create a second SPGP to

include wetlands in a different part of the State if it wishes

authority over a greater number of acres. Even a State with
62

an assumption program can create an SPGP to streamline

routine permits with minimal impacts. New Jersey attempted

to create such a general permit for the formation of

cranberry bogs in wetlands. Ultimately, the general permit

was rejected for lack of safeguards to pinelands in New

Jersey, but the option remains even under a State-assumed

permitting program.

Federal Oversight

The Corps views both the SPGP and State assumption as

successful programs because they lower the workload of the

Corps. Developers approve of such programs because they

believe that passing authority to States may be the only way

to reel in the numerous Federal regulatory arms. But,

SPGPs and State assumption do not automatically shift all

decisionmaking ability to the State level. Some of the control

a State might gain from creating an SPGP or assuming 404

permitting authority is tempered by the remaining Federal

oversight with these programs.

Under an SPGP, the Corp generally turns over to the State

the primary responsibility for individual permit review of the

activities included in the SPGP. This allows a State to

streamline permitting applications and approval for more

routine wetland disturbances but the EPA 404(b)(1)

guidelines provide that an SPGP “may be revoked or

modified by the Secretary of the Army if, after opportunity for

public hearing, the Secretary determines that the activities

authorized by such general permit have an adverse impact

on the environment or such activities are more appropriately

authorized by individual permits.” States also object to the

Corps’ authority to override State decisions on a case-bycase basis. The Corps retains authority to review PGP

applications individually and determine on a case-by-case

basis whether or not the concerns for the environment

require that the Corps override the permit and require an

individual application and review. This discretionary authority

is often incorporated as a condition to the programmatic

permit.

Similar Federal oversight exists under State assumption.

After assumption, the EPA monitors the effectiveness of the

State program on individual and overall levels. The State

must submit to the EPA a copy of each individual permit

application and proposed general permit. The EPA may

transfer permits to the Corps when deemed necessary. It

also requires annual reports to evaluate the State’s

administration of the program. State assumption documents

can also specify certain oversight authority with other

Federal agencies such as the Corps or Fish and Wildlife

Service. For instance, the MOA between New Jersey and

the Fish and Wildlife Service specified that activities

suspected of being in close proximity to sensitive areas will

be cleared by the Fish and Wildlife Service.

Even with these disadvantages of Federal oversight, the

State does gain more control. A State better controls the

timing of permitting decisions, the execution of onsite

evaluations, the drafting of permit conditions, and

establishes rapport with the public, which helps to maintain

public support for State wetland programs. Finally, the State

may be able to benefit from the mandatory Federal

oversight. When a State is faced with a tough political

decision, it can remove itself from the equation and shift the

decision to the Corps, leaving the Federal agency to take the

resulting “heat” from an unpopular permitting decision.

Flexibility

A final element of a State’s control over its regulatory

decisions under an SPGP or assumed program is flexibility.

Ideally, a State could experiment with parts of the program

until it best fits the State’s regulatory program already in

existence and the State’s wetland needs. Unfortunately, a

number of States cite lack of flexibility as a reason not to

partake in the SPGP or assumption alternatives.

Application procedures for the SPGP are less rigid than

assumption, and SPGPs provide a greater potential for

streamlined permit review. An SPGP can cover a limited

area in the State, such as the North Carolina SPGP that

covers 20 coastal counties, or can be statewide for particular

activities. The SPGP process also allows delegation of a

partial program to a State. The State may assume

management in the form of a pilot program on a small basis

before implementing the program across the State. This

allows a State to institute higher levels of regulatory

oversight for more fragile areas. For instance, Florida

recently developed an SPGP for a single watershed.

Funding

Congress relied on a State survey when providing for State

assumption in the CWA Amendments of 1977. The survey

revealed that States were interested in assumption but only

if adequate funding was available. Interestingly, Congress

has never provided such funding and does not appear to

consider it necessary. Lack of funds continues to be a top

deterrent for States considering assumption of the 404

program. The State agency must incur a large share of the

workload from the Corps including project review, impact

assessment, program enforcement and administration, and

the assumption of new responsibilities for compliance with

certain Federal statutes.

With an SPGP, States do not face the heavy funding burdens

associated with assumption of the section 404 program.

Because the SPGP is based on a State regulatory program

already in place, the State has already expended most of

the necessary funding. Its earlier investment in the

regulatory program will pay off when operating under an

SPGP. If additional funds are necessary and a State cannot

immediately fund the program, it can phase into operation

under an SPGP, lessening problems with start-up funding.

Or, a State can use an SPGP to strengthen an existing

regulatory program, using Federal funds to complement a

program limited by the shortage of State funds.

Permit Review

Both the SPGP and assumption alternatives offer a State a

way to change permit reviews and procedures and offer

more predictability in the permitting process. Because the

Corps is criticized for delayed permit review, States often

seek to simplify permitting procedures and streamline

application review and negotiations.

The SPGP expedites review of permit applications. An

SPGP can require accelerated action by those Federal
63

agencies that maintain veto or supervisory power. These

Federal agencies must meet State timelines that are often

shorter than the Federal counterparts. For example, the

North Carolina SPGP provides the Corps 45 days to

distribute the SPGP application materials to Federal

agencies to comment and develop a coordinated Federal

agency position. The SPGP allows for extension of this peer

review period, but it is discouraged. Similarly, the Maryland

SPGP offers a 45-day limit to the Corps but provides that the

Corps or the Maryland Department of Natural Resources

may extend the 45-day limit to complete the review. Finally,

in New Hampshire and Massachusetts, the Corps

coordinates with the Federal resource agencies through

meetings about every 3 weeks to review projects in the

SPGP screening category.

Operating under an SPGP also simplifies the permitting

process for applicants and avoids potential inconsistencies

in Federal and State approaches. Properly implemented, a

State can provide protection to wetlands while facilitating

minor development proposals. This bolsters the theory that

wetlands regulation is developing into a land use practice

and that the historical land use actor, the State, can best

define such policies.

Under State assumption, the State can design its permitting

procedures as long as it continues to meet EPA

requirements. States can limit delays in reviewing

applications and provide an avenue for better

communication between State regulators and permittees

and greater predictability in the application process.

Protection

Protection of wetlands remains a thorny issue in both State

and Federal regulatory programs. There is a constant

struggle to determine if the ability to provide adequate

protection for wetlands is at the Federal or the State level. If

a State decides to participate in regulation under an

assumed program or an SPGP, which alternative best

protects wetlands?

SPGP proponents cite increased protection of wetlands as a

benefit. Primary State control allows for increased

recognition of local differences in wetland conditions and

community needs. And it provides review to those smaller de

minimis impacts that are usually never reviewed. A

movement is gaining momentum to formulate an SPGP that

replaces Corps nationwide permits. This may be a result of

continuing criticism over the lack of examination of NWP

activities. Prior to the rise of SPGPs, in order to review

NWPs, States had to rely upon their authority to deny

section 401 water-quality certification to the Corps for an

NWP.

This approach supports delegating the NWP review to a

State under an SPGP because the Corps does not exercise

review over nationwide permit activities. This allows the

Corps to rely on a strong State program to take up the slack

and allows a State to protect against the dangers of

unregulated NWP activities. Under these circumstances, an

SPGP can provide a more streamlined review of minimal

impact activities that is better tailored to a State’s particular

circumstances.

The Corps’ New England Division is attempting this

approach. It has revoked several nationwide permits and

delegated those responsibilities to various States in its

region. The New Hampshire and Massachusetts SPGPs

have assumed many of these responsibilities but still ensure

at least some form of Federal review for impacts over 3,000

or 5,000 ft2, respectively. These may represent an

acceptable compromise between State protection and more

thorough Federal permit review.

Because the SPGP generally applies at the State

government level, the effectiveness of the program is limited

to that of the State agencies involved. Thus, SPGPs are

criticized for restricting local and regional involvement. In

addition, SPGPs may actually lower the amount of oversight

of wetlands by authorizing a State agency to substitute its

discretion in categorizing activities as those with minimal

impacts. Even small activities that have minimal impacts

cause loss of wetlands. While Corps oversight is present for

certain permit applications, the goal of reduction of

duplication and Corps workload may override the need for

adequate supervision. In testimony last year, a Corps

representative stated that “[p]rogrammatic general permits

allow the State or local agency to take the lead in working

with the applicant and reduce duplication among programs.

If enough of these programmatic general permits are

developed, the long-range benefit will be a significant

workload reduction for the Corps regulatory program.” The

primary goal, wetlands protection, is pushed aside in favor of

lessening duplication and lowering the workload for the

Corps.

State assumption has the potential to improve protection for

wetlands. States may offer more in-depth knowledge of local

and regional wetlands values and functions than Federal

agencies. In addition, a State may provide better

enforcement and may even create a program more stringent

than the Federal one. At a minimum, the State permits must

comply with the requirements of the CWA, its regulations

and the EPA 404(b)(1) guidelines, but this does not limit a

State as it may adopt more stringent requirements. For

example, a State may choose to require individual permits to

regulate isolated wetlands less than an acre in size, often

authorized under NWP 26 that would otherwise receive no

review. In 1984, the Detroit District of the Corps transferred a

majority of nationwide permits to the MDNR. The result is

that the Detroit district no longer issues NWPs in areas that

are under State jurisdiction in Michigan. There is no NWP

equivalent in Michigan, a significant change as a result of

the assumed program.

These benefits, however, may be outweighed by the

substantial political pressure a State agency can experience

when making permitting decisions previously made by the

Corps. Michigan and New Jersey have both experienced

some political challenges to their regulatory authority.

Michigan’s program has been cited as a success as the first

assumed program in the Nation. It has lowered the

workloads of both the EPA and the Corps, and both

agencies believe that the MDNR is doing an adequate job.

The Michigan program claims greater enforcement of

permitting violations than its Corps predecessor. However,
64

wetland protection advocates support Corps regulation and

cite great wetland losses due to the lack of political willpower

and enforcement mechanisms at the State level. One

commentator remarked that “in a large part because of its

permanence, the Michigan assumption is widely regarded by

conservationists to have been ultimately a disaster.”

Two examples can place this remark into context. First, the

EPA overruled its own regional office objection to a permit to

build a 267-acre gold course and housing development. In

1991, the Michigan DNR was set to issue a permit, but the

EPA regional office objected. After meetings between the

State, developer, and EPA regarding modifications, EPA

Administrator William Reilly withdrew the regional office’s

authority. An EPA representative explained that the action

was “to send a signal to the States that, if they follow the

proper procedures and are qualified and capable, EPA is

not going to interfere.” It did leave tension between the

MDNR, which was poised to issue a permit, and the

Regional EPA.

Second, the changes in the program in recent years have

caused strife within the program. Some cite Governor John

Engler’s guidance as weakening the program. In 1991,

Governor Engler issued Executive Order 1991–31, which

abolished some State agencies and reorganized others,

such as the MDNR. After litigation, the Michigan Supreme

Court upheld the restructuring in House Speaker v.

Governor, affirming the executive’s right to make such

changes. In 1994, the principal changes set forth in the 1991

Order were codified through statute, but, in 1995, Governor

Engler issued Executive Order 1995–18, further

reorganizing Michigan’s environmental agencies. The

National Wildlife Federation and State conservation

organizations challenged the 1991 and 1994 orders for

restructuring as violations of the State’s assumption

requirements, suing the EPA for failing to withdraw approval

of Michigan’s program due to substantial revision. The

District Court for the Western District of Michigan, however,

dismissed the claim, finding no subject matter jurisdiction.

The Michigan program remains in effect, but also in flux as a

result of the recent restructuring challenges. Its experience

shows that even at the State level, a wetlands regulatory

program may be subject to political whims. Also, it highlights

the hoops that a State must jump through to attain and

maintain an assumption program.

The New Jersey program has received favorable reviews

from the Federal agencies and many State conservationists

for its protection of wetlands. The program, however, faces

similar executive difficulties as the Michigan program. For

instance, conservationists in New Jersey have been vigilant

in opposing Governor Whitman’s attempts at creating a

general permit for cranberry bog conversions. Like the

struggles with reorganization of the Michigan program, New

Jersey’s program is also still subject to the impulses of State

political pressures and executive branches.

The experiences of these two States can assist other States

in determining if State assumption can help them protect

their wetlands.
65

LITTLE CYPRESS CREEK STUDY: A WATERSHED RESTORATION CASE STUDY

Lisa Gandy, Randy Roberson, and Tom Foti1

1

 Gandy, Project Scientist/Ecologist, FTN Associates, Ltd., 3 Innwood Circle, Suite 220, Little Rock, AR; Roberson, Natural Resources Planner,

AR Department of Parks and Tourism, One Capitol Mall, Little Rock, AR; Foti, Director of Research, AR Natural Heritage Commission, 323

Center Street, Little Rock, AR.

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability of

wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.

Abstract––The Little Cypress Creek watershed, which is the home of the Louisiana Purchase Historic State Park and

Natural Area, is one of the only remaining examples of a headwater swamp ecosystem left in Arkansas. An increase in

water elevations and a change in species composition were noticed in the park in late 1970. A study of the upper watershed

of Little Cypress Creek was conducted to identify potential factors causing these changes. Water elevations, vegetative

composition, and physical modifications were recorded and compared to historical information. A conceptual model of the

watershed was developed summarizing current understanding and hypotheses. Although natural variability in rainfall

explains some of the changes observed, roads, beaver dams, clearing, levee construction, and irrigation tail water inputs

contribute to hydrologic changes in the study area. The vegetation changes observed in the study area are likely due to

multiple stressors rather than any single factor.

INTRODUCTION

The Louisiana Purchase Historic State Park and Natural

Area (Park) is located in eastern Arkansas at the intersection of Monroe, Lee, and Phillips Counties (fig. 1). The 37-ac

Park contains a granite monument commemorating the 1803

purchase of the Louisiana Territory from France and marks

the point from which all land surveys of the Louisiana

Territory were referenced. The Park is not only historically

significant, but it also is ecologically significant in that it

preserves one of the remaining examples of a headwater

swamp ecosystem left in the State.

Arkansas Department of Parks and Tourism (Parks &

Tourism) and Arkansas Natural Heritage Commission

(Natural Heritage) personnel noticed that water levels were

increasing in the Park in the late 1970s and early 1980s.

Between an 18-to 24-in. increase in surface water elevations

was reported. In the early 1980s, Parks & Tourism and

Natural Heritage personnel observed that a large number of

overcup oaks (Quercus lyrata Walt.) on the slopes adjacent

to the deep swamp were dying; and species such as button

bush (Cephalanthus occidentalis L.) and smart weed

(Polygonum hydropiperoides Michx.) were becoming more

dominant.

The ability to sustain this unique headwater swamp

ecosystem and its rich heritage and values is an important

goal for Parks & Tourism and Natural Heritage. Achieving

this goal requires the cooperation of over 150 landowners

within the Little Cypress Creek watershed. In 1998, Parks &

Tourism and Natural Heritage initiated a study of the upper

watershed of Little Cypress Creek as the first step in the

process to achieve this goal. The objectives of this study

were: (1) to identify the factors that most likely have or are

contributing to changes in the upper watershed; and, (2) to

build community and stakeholder interest and ownership in

the restoration and long-term protection or sustainability of

the swamp.

The results of phase one of the Little Cypress Creek

watershed study are presented in this paper. The results of

the initial historical and current data analyses and the

community response to and involvement in the project are

shown. A hydrologic or water-balance model and model

analyses are being completed as phase 2, and the

development of a conceptual plan for restoration will be

completed as phase 3.

Figure 1—The Little Cypress Creek study area is located in

eastern Arkansas at the intersection of Monroe, Lee, and Phillips

Counties.
66

PROJECT APPROACH AND METHODS

Community Involvement

The ultimate goal of this project is to develop a communitybased restoration and protection plan that will be

implemented by landowners and other stakeholders within

the watershed. Community interest and involvement in the

study is a critical component in the overall success of this

project. Figure 2 shows the approach used to initiate and

foster community involvement in the project. The approach

consists of three interrelated and iterative steps: (1)

community and stakeholder outreach; (2) education and

community participation; and (3) transfer of project

ownership to stakeholders. Steps one and two were initiated

in this project.

Watershed Data and Analyses

The upper watershed of Little Cypress Creek was identified

as the limits of the study (fig. 3) to make best use of project

funding. As additional funding potentially becomes available,

data collection efforts will be expanded to the lower parts of

the watershed.

Available data and information describing historical and

current conditions in the study area were obtained from a

number of sources and through field surveys. These data

and information were compared to current data to identify

potential changes or factors within the study area that might

have impacted water levels and plant communities.

Drainage basin delineation––A preliminary delineation of

the study area basin was made based on 1971 USGS

topographic quadrangle maps (Monroe, Pine City, and

Marvell, ARK 7.5 minute series maps). Field surveys were

completed in April 1999 to confirm the preliminary drainage

area delineation. Due to the flatness of the area, the

drainage basin delineation will be refined as additional

information on drainage patterns in the study area are

collected.

Aerial photographs––Historical and current aerial

photographs were used to identify changes in land use,

drainage patterns, hydrologic modifications, and vegetation

within and adjacent to the study area. Several sources for

historical aerial photographs were investigated including the

U.S. Department of Agriculture (USDA) Farm Service

Agency (FSA) for Monroe and Phillips Counties, Arkansas

State Highway and Transportation Department, and the

USDA FSA Aerial Photography Field Office. The oldest aerial

photographs of the site were found at the USDA FSA county

offices. Black and white aerial photographs were obtained

for Phillips County for the period covering 1969 through

1972. Black and white photos dated 1982 were obtained for

Monroe County. No aerial photos were found for Monroe

County before 1982. No aerial photographs were needed for

Lee County because it was covered in the other

photographs.

Figure 2—A three-stepped approach is being used to initiate

and foster community involvement in the project.

Figure 3—The Little Cypress Creek study area is located within

the drainage basin of Big Creek within the White River watershed.

The upper drainage area of Little Cypress Creek was the limits of

phase 1 study area.
67

The USDA FSA Aerial Photograph Field Office obtained a

set of color infrared photographs of the study area dated

1992. The watershed was flown in May 1999 to obtain

current aerial photographs of the study area. Color infrared

aerial photographs were produced.

Hydrometeorological data––Historical and recent

precipitation and temperature data were available from the

National Weather Service monitoring stations at Brinkley,

Clarendon, and Marianna, AR. Monthly rainfall and

temperature data for these stations were downloaded from

CLIMATEDATAÔ (Hydrosphere 1998) for the period of

record covering January 1948 through December 1998.

Data for January 1999 through December 1999 were

obtained from the climatological reports issued by the

National Climatic Data Center. Daily rainfall and temperature

data for the study period, i.e., February 1999 through August

1999, were obtained from the local observers. Based on the

relative location to the study area, data from Clarendon and

Marianna were averaged to develop mean annual and

monthly precipitation totals for the study area. Data from

Brinkley were only used to fill in missing data from the other

two stations.

Runoff data––There are relatively few gauging stations

within the immediate vicinity of the study area. The closest

station is a discontinued station on Big Creek at Poplar

Grove (USGS station 07077950) located southeast of the

Little Cypress Creek study area. Data were available from

this station for the period from 1970 through 1993. These

data were supplemented with data from the L’Anguille River

at Palestine (USGS station 07047950) to develop a historical

record. This station was the closest station in the immediate

vicinity that had similar land use. A comparison of annual

runoff values from the L’Anguille River station with the Big

Creek at Poplar Grove station for the overlapping period

justified the use of the L’Anguille River data. The flow data

were expressed as runoff in inches per year. The annual

totals for the historical period and the mean monthly totals

were obtained.

Government Land Office record review––Government

Land Office survey records, housed at the Arkansas

Geologic Commission office, document early land surveys

throughout Arkansas. Records dating back to the original

establishment of the Baseline and the 5th Principle Meridian,

i.e., the location of the granite monument were reviewed.

Information on the historical composition of vegetation in the

watershed was obtained. Other vegetation accounts

specifically relating to the watershed are practically

nonexistent.

Arkansas Department of Heritage––Historical records

maintained at the Arkansas Department of Heritage’s

Arkansas History Commission and Natural Heritage

Commission offices also were reviewed to obtain available

information on Little Cypress Creek upper watershed and

the Louisiana Purchase Historic State Park. The available

records spanned a time frame over the past 100 years.

Arkansas Department of Parks and Tourism––A review of

files and records at Parks & Tourism’s Planning and

Development office was completed to obtain additional

information on the historical condition or on activities and

changes that have been documented for the Park.

Information in these files covered a 30-year time frame from

mid-1960 to mid-1990.

Water surface elevation measurements––No historical

water-surface elevations were available or are known from

the study area. To document current conditions within the

study area, water-surface elevation data were collected at

five locations within the study area (fig. 4). Three types of

recording devices were used to measure water-surface

elevations within the watershed: continuous stage recorders,

staff gauges, and crest gauges.

Continuous stage recorders––StevensTM stage recorders

were placed at three locations within the Little Cypress

Creek study area (fig. 4) to record water level data on a

continuous basis during the study. One recorder (recorder 7)

was placed adjacent to the granite monument to record

water levels within the Park. A second recorder was placed

on the upstream side of the road at the Cotton Trailer Road

(recorder 3). This road is located in the approximate middle

of the study area. The third recorder (recorder 9) was placed

in the lower portion of the study area on the downstream

side of Rogers Road. The monument and Cotton Trailer

Road recorders were installed in the watershed in February

1999. The third recorder at Rogers Road was installed in

June 1999 after a review of water-surface elevation data

indicated that a recorder was needed in the lower part of the

study area.

Figure 4—Locations of road crossings and water-surface elevation

recorders in the study area.
68

The stage recorders were set to record changes in watersurface elevations over a 16-day period. Charts were

replaced every 2 weeks from February 4, 1999, through

August 9, 1999. Electronic data loggers were initially placed

on the recorders at the monument and at the Cotton Trailer

Road. Continuous water-elevation data were captured with

the data loggers for a period of approximately 1 to 2 months,

and the data were cross-referenced to the chart data to

verify the operation of the recorders.

Staff gauges––A total of nine staff gauges were installed at

five locations throughout the study area (fig. 4) to obtain

instantaneous water-surface elevations. Only one staff

gauge (staff gauge 7) was placed in the Park near the

granite monument. Staff gauges were placed on the

upstream and downstream sides of the North Road (staff

gauges 1 and 2, respectively), the Cotton Trailer Road (staff

gauges 3 and 4, respectively), the South Road (staff gauges

5 and 6, respectively), and Rogers Road (gauges 8 and 9,

respectively). Staff gauges were surveyed at the beginning

and end of the study. Staff-gauge readings were taken on a

biweekly basis from February 4, 1999, to August 9, 1999.

Crest gauges––A total of six crest gauges were installed at

three locations throughout the study area (fig. 4) to obtain

the maximum surface-water elevation at designated

locations within a 2-week period. Readings were taken on a

biweekly basis from February 4, 1999, through August 9,

1999. Crest gauges were surveyed at the beginning and end

of the study period.

Spot elevations––No historical survey data documenting

land-surface elevations within the headwater swamp or

elevation of the granite monument were found. The

elevations of the recorders and gauges, top of the granite

monument in the Park, and the ordinary high-water mark on

the granite marker were surveyed. Road profiles were shot

for the North Road, Cotton Trailer Road, South Road, and

Rogers Road to determine the elevation of the lowest point

along each road. Where culverts were present under the

road, the elevation was obtained for the invert of each

culvert.

Surface water withdrawals and irrigation––The Arkansas

Soil and Water Conservation Commission was contacted to

identify whether any registered riparian or nonriparian water

users were located within the study area. Registered users

are persons that withdraw greater than or equal to 1 ac ft of

water per year from a stream or other waterbody within the

State. The USDA Natural Resources Conservation Service

County offices in Monroe and Phillips Counties were

contacted to identify any water withdrawals within the study

area.

Groundwater withdrawals––The U.S. Geologic Survey

(USGS) was contacted to obtain a list of irrigation or other

wells within the watershed. Although there are several

irrigation wells in the study area (USGS 1968, 1984), no

information on the amount of groundwater withdrawn from

the study area was available from the USGS.

Vegetation surveys––Vegetation surveys were completed

in the Little Cypress Creek study area on March 30, April 13,

and May 13, 1999, to make observations of existing

vegetation. These surveys also were completed to identify

potential reference areas within the watershed for

restoration. Dominant species, community types, and

condition of vegetation were observed and recorded.

RESULTS

Aerial Photograph Review

Land use changes—Approximately 65 percent of the

approximately 7,680-ac study area, i.e., approximately 12

mi2, is currently in agricultural crop production with the

remaining 45 percent in forests. Historically, hardwood

forests were extensive within the watershed. The majority of

land clearing in the study area occurred prior to 1969. Land

clearing has been minimal in the Little Cypress Creek study

area since that time. A total of 440 ac were cleared in the

lower half of the study area, i.e., below the Park, between

1969 and 1982. No additional land clearing of significance

was identified between 1982 and 1999. Overall, < 10 percent

of the land within the study area had been converted from

forest to agricultural land use between 1969 and 1999.

Drainage—The comparison of historical and current aerial

photographs did not reveal any obvious changes in drainage

patterns within the study area from 1969 to the present.

However, discussions with landowners indicated that large

changes in drainage patterns have occurred within the

watershed, including an increase in pumped drainage and

rerouting of flows within the watershed. Detailed mapping of

drainage patterns was not conducted as part of this phase of

the project but general patterns were recorded. Fields on the

far-west side of the study area generally pump water into

ditches that ultimately drain to the lower part of the study

area. Most of the fields close to or adjacent to Little Cypress

Creek either pump or discharge directly into the creek.

Several ditches along the west and east sides of the study

area carry water directly into Little Cypress Creek.

Levees—Levees have been constructed throughout the

watershed to eliminate inundation of adjacent farmland or to

hold water during the winter for hunting leases. Landowners

in the study area indicate that since the early 1950s, a

significant number of acres that historically provided storage

of water within the system have been cleared and leveed.

Approximately 320 ac were cleared and leveed between

1979 and 1982. Without additional historical and current

information, the reduction in storage capacity within the

study area could not be adequately quantified, but it is

estimated that approximately one-third of the historical

storage capacity of the study area has been lost.

Irrigation—Based on conversations with landowners, there

has been an increase in the acres of irrigated lands within

the study area since the early 1980s due to the conversion

from soybeans to rice production. Within the study area,

groundwater is continuously pumped onto the fields to

eliminate stagnation of water and increased incidence of

pests. Consequently, irrigation tail water is continuously

discharged into Little Cypress Creek from April through

September. Rapid discharges of large volumes of water into

Little Cypress Creek also occur at the end of the ricegrowing season in preparation for harvest. The volume of
69

water entering the system from irrigated fields was not

quantified as part of this phase of the study.

Roads—The historical aerial photographs showed that a

number of road crossings existed within the study area prior

to 1969. Figure 4 shows the locations of three major

crossings south of the Park, i.e., Cotton Trailer Road, South

Road, and Rogers Road, and one north of the Park, i.e.,

North Road, that were present in the study area in 1969 and

that currently exist. No new road crossings were constructed

in the watershed since 1969. The roadbed of Cotton Trailer

Road was elevated some time in the 1990s to its present

height.

Hydrometeorological Data

Precipitation––Annual precipitation totals in the study area

range from 35.3 to 71.5 in. over the approximately 42-year

period of record (table 1), and the long-term mean annual

precipitation total is 51.1 in. A comparison of rainfall data for

the study period, i.e., 1999, to historical data indicates that

total precipitation for 1999 was approximately 4 in. less than

the long-term average. The highest total annual rainfall of

71.5 in. was observed in 1979 and was approximately 20 in.

greater than the mean annual total rainfall. Observations of

an approximate 18-to 24-in. increase in water elevations in

the Park made by Parks & Tourism and Natural Heritage

personnel in the late 1970s are consistent with this data.

Average monthly precipitation totals for the study area (fig. 5)

for the period of record range from approximately 3.0 to

5.3 in. The greatest amounts of precipitation typically fall

during the months of November through May. The driest

months are June through October. During the study period,

conditions were much drier than the average during the fall

of 1998. With the exception of February 1999, conditions

during the study period were wetter in the winter and spring

than long-term monthly mean values.

Runoff––The delineation for the upper Little Cypress Creek

drainage area was based on historical topographic maps

and a ground reconnaissance. Due to the flatness of the

area and the many anthropogenic modifications for

drainage, the basin boundaries are difficult to define and

probably have changed over the years. Runoff averaged

18.3 in. per year or approximately 36 percent of precipitation.

On a monthly basis, runoff varied from 0.46 in. or 12 percent

of precipitation in July to 2.54 in. or 58 percent of

precipitation in February.

Groundwater

Although irrigation wells exist within the study area (USGS

1968, 1984), there is no published groundwater level data to

characterize groundwater levels within the Little Cypress

Creek watershed or to confirm whether groundwater

discharges to the Little Cypress Creek wetland ecosystem.

Papers describing the geomorphology of the Little Cypress

Creek drainage area and the HGM classification of wetlands

in Arkansas (Klimas 1999, Saucier 1996) indicate that there

is the potential for groundwater discharge into Little Cypress

Creek. Based on groundwater data (USGS 1998) collected

from a well located less than one-fourth of a mile south of

the study area (local well 1S01E20DDB1), the water table

may be as close as 10 to 20 ft below ground surface at

ground surface elevations of 185 ft.

Hydrographs from wells located throughout the region show

a general decrease in groundwater levels, since the mid1950s (USGS 1998). Landowners with irrigation wells within

the study area indicated that groundwater within the study

area has not exhibited this trend. These observations are

consistent with groundwater elevation data collected from a

local well several miles south of the study area in Phillips

County (local well 02S01E28CCB1). Data collected since

1955 from this well show that groundwater levels have

varied over the 40-year period of record but have not been

decreasing.

Water-Surface Elevations

Time series plots of water-surface elevations at the

monument and at Cotton Trailer Road indicate that watersurface elevations within the wetland ecosystem follow

seasonal rainfall patterns, increasing during the winter and

early spring, then dropping during late summer and early

fall. Early in the season, water levels in the swamp also rise

in response to individual storm events (fig. 6). The

corresponding rise in surface-water elevations with large

storm events occurs mostly before leaf out and disappears

during the latter part of the growing season when

evapotranspiration is at its highest. The measured increase

in water-surface elevations is greater than total precipitation

inputs into the basin, indicating that runoff is an important

hydrologic input into the system.

During the study period, maximum water-surface elevations

in the Park were approximately 10 in. lower than the highwater mark on the granite marker. The high water levels

observed in the late 1970s and 1990s were not observed

during this study. Staff gage readings taken on the upstream

and downstream sides of four major roads show that

surface-water elevations on the upstream sides of the roads

typically were higher than the downstream sides.

Figure 5—Comparison of monthly rainfall totals during the study

period with long-term monthly average rainfall.
70

Surface-water profiles of the study area for the months of

March, June, and July 1999 provide a graphical depiction of

surface-water elevations relative to location in the watershed

and the roads (figs. 7, 8, 9). Water-surface elevations drop

over the season and typically are higher on the upstream

side than the downstream side of the road. The profile from

July shows an increase in surface-water elevations between

the South Road and Rogers Road during a period when the

water levels are dropping in other parts of the ecosystem.

Landowners indicate that at times, water flows backwards,

i.e., upstream, over the South Road.

Vegetation Assessment

The original survey notes of the Louisiana Purchase

Monument area provide strong evidence that the basic

vegetation and landscape features immediately surrounding

the granite monument probably have changed little since

1815. The survey notes do not adequately distinguish

between certain critical species. However, the records

clearly denote that areas presently occupied by tupelo

swamp were occupied by swamp in 1815. From the historical

accounts, it appears that the area surrounding the granite

monument ranged from being flooded to dry during the

summer months. No historical information was found

describing the vegetation in other areas of the Park or within

the study area.

The upper watershed of Little Cypress Creek contains a

diversity of upland and wetland plant communities. The

central core of the wetland ecosystem supports the deepest

water levels, i.e., 2 to 5 ft deep, and is dominated by water

tupelo (Nyssa aquatica L.), although, in some cases, bald

cypress (Taxodium distichum L.) also occurs. Historically,

bald cypress was codominant in the deep swamps. With the

exception of a few depressions within the deep swamp,

timber harvest activities have removed most of the bald

cypress from this community.

The shallowly flooded areas on the west side of the Park

were dominated by overcup oak, red maple [Acer rubrum L.

var. rubrum (BB) and A. rubrum L. var. drummonii (H. & A.)

Sarg.], sweet gum (Liquidambar styraciflua L.), and button

bush. Several dead snags of overcup oak are present within

this community, and those that are alive exhibit morphological features indicative of stress. Tree branches are short,

thickened, and knarled, and many exhibit epicormic

branching. Almost all of the dead trees appear to be overcup

oak and approximately the same age. The 1999 field surveys

also identified that many overcup oaks near the entrance of

the Park appear to be infected with a canker rot fungus. A

large percentage of the oaks in this part of the Park and

along the road leading into the Park are heavily infested with

insect galls.

Beaver Impacts

Beaver dams exist throughout the watershed. Large beaver

dams were recorded at the North Road, at the Cotton Trailer

Road, in the vicinity of the South Road, and at the outlet end

of the study area where Little Cypress Creek crosses

Arkansas State Highway 49. Other beaver dams likely exist

within the study area. Most of the beaver activity appears to

take place in the more shallowly flooded areas and where

there has been disturbance. The beaver-impacted area

upstream of the North Road appears to have been in

existence for approximately 30 years based on the size and

appearance of the black willow (Salix nigra Marsh.) trees.

Some of these areas also appear to have been cut over

repeatedly by beaver. Beaver activity has resulted in a

degradation of bottomland hardwood forests and loss of

overstory and resulted in increased scrub/shrub

communities dominated by black willow and buttonbush.

Timber Harvest

Historical and recent timber-harvesting activities have

occurred throughout the study area. In drier portions of the

swamp such as the area adjacent to the Park on the east,

there is ample evidence of former timber-harvest activities in

the form of stumps and downed logs. An open area within

the swamp just upstream of the South Road has been

impacted by timber harvest of bald cypress. No pole-sized

timber or seedlings have become reestablished in this area.

DISCUSSION

Over the past decade there has been a shift away from the

management and protection of natural resources on a siteby-site basis to the management and protection of natural

resources on an ecosystem basis, i.e., watershed,

landscape, or regional scale (Sparks 1995). The

assessment, management, and protection of resources is

being conducted on spatial scales ranging from millions of

acres, such as in the Chesapeake Bay and the Florida

Everglades, to hundreds of acres in individual drainage

basins and watersheds. Current programs and regulatory

mandates, e.g., unified watershed assessments, total

maximum daily load studies, etc., are evidence that policy

and decisionmakers, natural resource managers, scientists,

regulators, and the public recognize that the sustainability of

natural resources is dependent on the management of these

resources at larger scales.

The attempt to manage resources and water within the Little

Cypress Creek study area on an individual-by-individual

farm basis has failed. As landowners have cleared, drained,

Figure 6—Continuous stage recorded within the Park (recorder 7),

and daily rainfall from January 1999 through August 1999.
71

Figure 7—Water-surface profile in the Little Cypress Creek upper watershed

study area on March 16, 1999.

Figure 8—Water-surface profile in the Little Cypress Creek upper watershed

study area on June 30, 1999.

Figure 9—Water-surface profile in the Little Cypress Creek upper watershed

study area on July 23, 1999.
72

or held water on one property, others have been flooded and

lost timber on other parcels. Although small relative to other

large-scale restoration and management efforts in the

United States., a comprehensive watershed approach is

needed to manage, restore, and sustain the values and

condition of the Little Cypress Creek headwater swamp

ecosystem.

Because little historical data is available to compare to

current data and conditions, it is difficult to clearly identify

and quantify the changes within the headwater swamp

ecosystem. In the absence of historical data, the data

collected in the phase 1 study were summarized in the

context of a conceptual model (fig. 10). This conceptual

model focuses on the changes in hydrology within the upper

watershed, which can then be linked or correlated to other

changes within the watershed, e.g., changes in plant

community composition, changes in wildlife habitat and

usage, etc. This conceptual model summarizes the

hypotheses and assumptions that can be tested or

evaluated and serves as a framework for identifying data

gaps and future data needs or studies. The conceptual

model will be modified and expanded over time as new

information and data are gathered, the natural variability

within the system is better quantified, and as changes or

perturbations continue to occur within the system.

Changes in Storage

Alterations in the Little Cypress Creek upper watershed have

had an influence on the historical hydrologic regime of the

headwater swamp ecosystem. These changes can be

expressed as a change in the water balance within the

system. Although alterations within the watershed can either

increase or decrease water stored within the swamp

ecosystem, the direction of change in the Little Cypress

Creek study area has been weighted more toward an

increase in the amount of water stored within the Little

Cypress Creek headwater swamp ecosystem (fig. 10).

Like most watersheds in the Mississippi Alluvial Plain of the

Lower Mississippi Valley, bottomland and upland forests of

the Little Cypress Creek watershed have been cleared for

agricultural production. Approximately 65 percent of the Little

Cypress Creek study area has been cleared of its original

forest cover and is in agricultural production. Clearing within

the Little Cypress Creek study area likely has changed the

amount and timing of runoff into the Little Cypress Creek

headwater swamp. Continuous water-surface elevation data

collected at the stage recorders indicate that changes in

stage are greater than the total rainfall of any single storm

event. The data suggest surface runoff during a storm event

is an important input into the system.

Crop changes from soybeans to rice also have changed the

hydrology of the Little Cypress Creek ecosystem. The timing

of hydrological inputs has been altered, and inputs of water

into the system are currently being extended into late

summer and fall when natural precipitation and runoff inputs

typically begin to decrease. Groundwater has now become a

surface-water input into the system as it is pumped onto

fields and discharged into the creek on a continuous basis

from April through September.

Levee construction within the study area has largely been

successful in keeping water off many parcels that historically

were wet and supported bottomland hardwood forests.

Levee construction, however, has effectively reduced the

historical storage capacity, constricted drainage, and

reduced total storage area within the watershed. Belt (1977)

and Myers and White (1993) have documented that the

constriction of the floodway has increased flood stage on the

Mississippi River. Although the hydraulics of the Little

Cypress Creek system are very different from the

Mississippi River, the loss of storage capacity and

constriction of drainage by levee construction can have the

same effect. Without a concomitant reduction in hydrologic

inputs, the amount of water stored within the remaining

headwater swamp ecosystem and water levels will increase

on a per-acre basis.

Road construction and the introduction of beaver have also

contributed to the alteration of the natural hydrology of the

system. The roads and the beaver dam at the North Road

have compartmentalized the ecosystem and impeded

natural flows within and discharge out of the system. In this

low gradient, flat ecosystem where there is < 1 ft difference

in elevation from the top of the watershed to the outlet end,

water is easily impounded on the upstream sides of the

roads increasing both the aerial extent of inundation and

water depths within each compartment. The total acres

inundated and depth of water within each compartment will

depend on the amount, timing, and duration of the various

hydrologic inputs into each of these compartments.

Above normal rainfall totals and increased runoff that

occurred in the late 1970s and in the early 1990s are

consistent with increased storage and increased surfacewater elevations observed in the Park since the late 1970s.

Based on these data alone, it would suggest that some of

the increases in surface-water levels observed in the Park in

the late 1970s through the early 1990s were due to natural

variability in rainfall. Although no data are available to

document whether water storage has increased in the Little

Cypress Creek headwater swamp, it is indicated by this

study that the cumulative effects of increased runoff, loss of

storage capacity, constriction and impediments in the

drainageway, and new hydrologic inputs should increase the

Figure 10—Conceptual model of potential factors affecting

water storage within the Little Cypress Creek upper

watershed study area.
73

amount of storage within the remaining headwater swamp

ecosystem.

Vegetation Changes

Beaver impacts have become widespread throughout

Arkansas since the reintroduction of this species in the State

in the 1950s (Selander 1979). Beaver activities within the

study area have led to a marked increase in sedimentation,

loss of forest cover, and changes in vegetative composition.

Some of these areas appear to have been affected by

beaver for at least 30 years.

Changes in the duration, depth, and timing of flooding or

inundation of bottomland hardwood systems are

documented to have an effect on the survival of many

bottomland hardwood trees (Theriot 1993, Whitlow and

Harris 1979). Although it is reported that major changes in

water level may cause mortality in water tupelo (Johnson

and Beaufait 1965), no evidence of dead or stressed tupelo

was observed within the Park or in the rest of the study area.

There is evidence of mortality in overcup oak near the

entrance of the Louisiana Purchase Historic State Park. The

oak mortality at the entrance to the Park appears, however,

to represent an episodic event rather than a continuing

occurrence at the site. Almost all of the dead trees appear to

be overcup oak, and most appear to be of approximately the

same age. In addition to mortality, many of the overcup oak

trees that remain in this part of the Park are infected with

what appears to be a canker rot fungus. Canker rot fungus

has been reported to occur in bottomland hardwood forests

in the South and to cause mortality in many wetland species

(McCracken 1978), but overcup oak is not reported to be

susceptible to canker rot (Morris 1965, Solomon 1990).

The reported flood tolerance of overcup oak varies from

being very tolerant of deep flooding, i.e., flooding for more

than 1 year (Theriot 1993, Whitlow and Harris 1979) to

moderately tolerant; i.e., able to tolerate saturated or flooded

soils for several months during the growing season but

susceptible to high mortality if flooding persists or reoccurs

several consecutive years (Hook 1984, McKnight and others

1981). A few reports indicate that high mortality in overcup

oak is likely if flooding persists or reoccurs for a period

greater than 3 years (Teskey and Hinckley 1978).

Historical precipitation data, however, show an approximate

10-year period of below average rainfall conditions from

1962 to 1972. The study suggests that dry conditions in the

ecosystem from 1962 to 1972 could be as important a factor

in stressing overcup oak as increased storage of water in

the late 1970s. Periodic occurrences of oak decline have

been reported in many areas of the southeast (Wargo and

others 1983). The causes of oak decline are very complex

and often are due to the interaction of multiple stressors

(Leininger 1998, McCracken and others 1991, Wargo and

others 1983). A combination of drought, flooding, changes in

hydroperiods, and pest infestation could be important factors

in the decline of overcup oak-dominated forest in the Park.

Community Involvement

Initial public participation in and response to this study has

been positive. Many landowners cooperated with the study,

and a partnership with the local high school was formed.

Over 40 individual stakeholders have provided support to

the project in the form of interest in the project, information,

permission to enter property, permission to install scientific

equipment on private property, and collection of data.

Involved landowners and stakeholders would like to see the

Little Cypress Creek watershed restored to: (1) reduce

flooding on their properties; (2) increase ability to farm

marginal property; (3) restore lost bottomland hardwoods to

increase income from hunting leases; and, (4) restore

bottomland hardwood values for personal recreational

opportunities. With a core group of landowners showing

interest in the management and restoration of the Little

Cypress Creek ecosystem, the initial steps in gaining

community involvement in the project were considered

successful. Like other community-based restoration

programs that have been developed throughout the country

to sustain natural resources (EPA 1997, French 1999), it is

anticipated that it will take years to build a community-based

alliance focused on the restoration of the Little Cypress

Creek watershed.

The information collected in this first phase of the Little

Cypress Creek watershed study has only begun to scratch

the surface of the information needed to assess condition,

determine restoration alternatives, and develop

management plans for the watershed. Nevertheless, this

study of the Little Cypress Creek upper watershed has lead

to a number of important findings relative to the project goals

and objectives; and community awareness of the issue has

increased. A more organized or formal vehicle for

stakeholder involvement has been established as a result of

the study along with identification of those areas that are

critical for continued investigation.

ACKNOWLEDGMENTS

The Little Cypress Creek Upper Watershed Hydrologic and

Environmental Study was funded through the U.S.

Environmental Protection Agency Region 6, State Wetlands

Grants program. The Arkansas Department of Parks and

Tourism and the Arkansas Natural Heritage Commission

provided matching funds for the project. The authors wish to

acknowledge the following individuals who contributed

greatly to the project: Dr. Dennis E. Ford, Shannon P.

Holbrook, Dr. Gary E. Tucker, and Gerry T. Conley of FTN

Associates, Ltd.; Brinkley High School East lab students;

and the Little Cypress Creek landowners.

REFERENCES

Belt, C.B. 1977. The 1973 flood and man’s constriction of the

Mississippi River. Science. 189: 681–684.

French, L.J. 1999. A case study in forming a grassroots watershed

organization. DE: Nanticoke Watershed Alliance. 3 p.

Hook, D.D. 1984. Waterlogging tolerance of lowland tree species in

the South. Southern Journal of Applied Forestry. 3: 136–149.

Hydrosphere [Corporation]. 1998. CLIMATEDATATM CD-ROM,

NCDC summary of the Day-West2. Boulder, CO: Hydrosphere

Inc. Vol. 9.3.
74

Johnson, R.L.; Beaufait, W.R. 1965. Water tupelo (Nyssa aquatica

L.). In: Fowells, H.A. comp. Silvics of forest trees of the United

States. Agric. Handb. 271. Washington, DC: U.S. Department of

Agriculture: 284–286.

Klimas, C. 1998. Preliminary classification of wetlands in Arkansas

based on hydrogeomorphic criteria. Little Rock, AR: Arkansas

Multiagency Wetland Planning Team.

Leininger, T.D. 1998. Effects of temperature and drought stress on

physiological processes associated with oak decline. In: Mickler,

R.; Fox, T., eds. The productivity and sustainability of southern

forest ecosystems in a changing environment. New York:

Springer Verlag, Inc.: 647–661.

McCracken, F.I. 1978. Canker rots in southern hardwoods. Forest

Insect and Disease Leafl. 33. Stoneville, MS: U.S. Department of

Agriculture, Forest Service, Southern Forest Experiment Station.

4 p. In cooperation with: Mississippi Agricultural and Forestry

Experiment Station; Southern Hardwood Forest Research

Group. [Internet version].

McCracken, F.I.; Ammon, V.; Solomon, J.D.; Nebeker, T.E. 1991. Oak

decline in the Lower Mississippi River Valley. In: Coleman, S.S.,

Neary, D.G., comps. Proceedings of the sixth biennial southern

silvicultural research conference; 1990 October 30–November 1;

Memphis, TN. Gen. Tech. Rep. 70. Asheville, NC: U.S. Department of Agriculture, Forest Service, Southeastern Forest

Experiment Station: 299–306. Vol. 1.

McKnight, J.S.; Hook, D.D.; Langdon, O.G.; Johnson R.L. 1981.

Flood tolerance and related characteristics of trees of bottomland forests of the Southern United States. In: Clark, J.R.;

Benforado, J., eds. Wetlands of bottomland hardwood forests.

New York: Elsevier Press: 26–69.

Morris, R.C. 1965. Overcup oak (Quercus lyrata Walt.). In: Fowells,

H.A., comp. Silvics of forest trees of the United States. Agric.

Handb. 271. Washington, DC: U.S. Department of Agriculture:

600–602.

Myers, M.A.; White, G.E. 1993. The challenge of the Mississippi

flood. Environment. 10(12): 6–35.

Saucier, R.T. 1994. Geomorphology and quarternary geologic

history of the Lower Mississippi Valley. Vicksburg, MS: U.S. Army

Engineers Waterways Experiment Station. 364 p. Vol. 1.

Selander, J. 1979. A guide to Arkansas mammals. Conway, AR:

River Road Press. 313 p.

Solomon, J.D. 1990. Overcup oak (Quercus lyrata Walt.). In: Burns,

R.M.; Honkala, B.H., tech. coords. Silvics of North America.

Agric. Handb. 654. Washington, DC: U.S. Department of

Agriculture, Forest Service. 877 p. Vol. 2.

Sparks, R.E. 1995. Need for ecosystem management of large rivers

and their floodplains. BioScience. 45(3): 168–182.

Teskey, R.O.; Hinckley, T.M. 1978. Impact of water level changes on

woody riparian and wetland communities. southern forest region.

FWS/OBS–78/87. Columbia, MO: U.S. Department of the

Interior. U.S. Fish and Wildlife Service, Office of Biological

Services. 46 p. Vol.2.

Theriot, R.F. 1993. Flood tolerance of plant species in bottomland

forests of the Southeastern United States. Wetlands Res.

Program Tech. Rep. WRP–DE–6. Vicksburg, MS: U.S. Army

Engineers Waterways Experiment Station, Environmental

Laboratory. 50 p.

U.S. Environmental Protection Agency. 1997. Top 10 watershed

lessons learned. EPA/840–F–97–001. Washington, DC: U.S.

Environmental Protection Agency, Office of Water, Office of

Wetlands, Oceans, and Watersheds. 59 p.

U.S. Geological Survey. 1968. Quaternary aquifers in the Mississippi

Embayment. Geol. Sur. Prof. Pap. 448–E. Washington, DC: U.S.

Government Printing Office. 15 p.

U.S. Geological Survey. 1984. Maps showing altitude of the

potentiometric surface and changes in water levels of the alluvial

aquifer in eastern Arkansas, spring 1983. Water Res. Invest.

Rep. 84–4264. Little Rock, AR: U.S. Department of the Interior,

U.S. Geological Survey. 1 p. In cooperation with: Arkansas

Geological Commission and the Arkansas Soil and Water

Conservation Commission.

U.S. Geological Survey. 1998. Status of water levels and selected

water quality conditions in the Mississippi River Valley alluvial

aquifer in eastern Arkansas, 1994–96. Water–Res. Invest. Rep.

98–4131. Little Rock, AR: U.S. Department of the Interior, U.S.

Geologic Survey. 72 p. In cooperation with: Arkansas Soil and

Water Conservation Commission and the Arkansas Geological

Commission.

Wargo, P.M.; Houston, D.R.; LaMadeleine, L.A. 1983. Oak decline.

Forest Insect and Disease Leafl. 65. Hamden, CT: U.S. Department of Agriculture, Forest Service, Northeastern Forest

Experiment Station; Northeastern Area State and Private

Forestry. 6 p. [Internet version].

Whitlow, T.H.; Harris, R.W. 1979. Flood tolerance in plants: a stateof-the-art review. Tech. Rep. E–79–2. Vicksburg, MS: U.S. Army

Engineers Waterways Experiment Station, Environmental

Laboratory. 161 p.
75

INTRODUCTION

Afforestation activities in the Lower Mississippi River Alluvial

Valley (LMRAV) are currently peaking after the past four

decades of extensive deforestation (Allen 1997, King and

Keeland 1999, Stanturf and others 1998). Different forces

drive the afforestation efforts of public resource agencies,

conservation organizations, private corporations, and private

landowners. Their interests include conversion of

economically marginal agricultural land to forest cover for

ecosystem rehabilitation, soil conservation, aesthetics, and

recreation. They may want to establish intensively managed

fiber farms and carbon sequestration banks, and to mitigate

forested wetlands destroyed elsewhere in development

projects.

Since 1992 on more than 250,000 ac in Louisiana,

Mississippi, and Arkansas, governmental incentive programs

have defrayed plantation establishment costs and/or

purchased long-term conservation easements from private

landowners (King and Keeland 1999, Stanturf and others

1998). Since program dollars largely support these

afforestation activities, associated enrollment deadlines

often create a sense of urgency about plantation

establishment. The urgency with which managers have

implemented recent afforestation activities leads some to

question the validity of afforestation decisions and practices

at the stand, forest, and landscape levels (Allen 1997;

Stanturf and others, in press; Twedt and Portwood 1997;

Wilson and others, in press).

A sustained, formal research initiative on bottomland

hardwood plantation establishment in the LMRAV began as

early as the 1960s (Kennedy 1993). This manuscript

summarizes current afforestation techniques and compares

the knowledge base with current practices to emphasize

aspects of afforestation where existing knowledge is not

incorporated into operational practice. This manuscript also

reviews new silvicultural systems to enhance success on

adverse sites and enhance ecological benefits, while

addressing multiple objectives of landowners. In addition to

available literature, a large portion of this manuscript is

based on the authors’ combined observations and

experiences in the LMRAV.

REVIEW OF CURRENT AFFORESTATION

TECHNIQUES

Management Objectives

Defining clear management objectives in terms of what

outputs are to be achieved is prerequisite to directing

successful establishment and future management of

afforestation sites. Landowner objectives will ultimately

influence or govern all system decisions involving stand

establishment (species suitability, site preparation

requirements, planting density, postplanting operations),

intermediate stand management (precommercial thinning,

timber stand improvement, stand health and sanitation

practices, improvement cutting and thinning, control of fire,

disease, insects, or other damaging agents), and

regeneration harvesting (including regeneration methods) in

order to regulate forest structure for the desired outputs

(Daniel and others 1979). Management objectives can be

simple or complex encompassing multiple aspects of

watershed or wildlife management, wood, fiber or forage

production, and aesthetics (Daniel and others 1979, Smith

1986). Rehabilitating functions and health of the bottomland

hardwood forest ecosystem always underlies other

management objectives.

BOTTOMLAND HARDWOOD AFFORESTATION: STATE OF THE ART

Emile S. Gardiner, D. Ramsey Russell, Mike Oliver, and Lamar C. Dorris, Jr.1

1 Gardiner, Research Forester, Center for Bottomland Hardwoods Research, Southern Research Station, USDA Forest Service, Stoneville, MS

38776; Russell, Administrative Forester, North Mississippi Refuges Complex, US Fish and Wildlife Service, Grenada, MS 38901; Oliver, Project

Biologist, Ducks Unlimited Inc., Vicksburg, MS 39180; Dorris, Administrative Forester, Yazoo National Wildlife Refuge Complex, US Fish and

Wildlife Service, Hollandale, MS 38748.

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability of

wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.

Abstract—Over the past decade, land managers have implemented large-scale afforestation operations across the

Southern United States to rehabilitate agricultural land historically converted from bottomland hardwood forest cover types.

These afforestation efforts were initially concentrated on public land managed by State or Federal Government agencies,

but have later shifted towards private holdings that qualified for governmental assistance or cost-share programs. Traditional

silvicultural practices dominate bottomland hardwood afforestation schemes in the South, with 1-0 bare-root oak (Quercus

spp.) seedlings comprising the balance of planting stock mixtures. However, traditional methods do not always yield

successful afforestation, especially when applied on an operational scale across a landscape of heterogeneous site types

and ownership objectives. This manuscript summarizes bottomland hardwood afforestation techniques and compares the

knowledge base with current practices. Additionally, this manuscript reviews new silvicultural systems to enhance establishment success on adverse sites, to enhance ecological benefits of afforestation, and to address multiple objectives of

landowners. We identify four vital components of afforestation that are generally lacking in most regeneration activities in the

Lower Mississippi River Alluvial Valley.
76

Government monetary incentives for landowners to establish

forest cover drive objectives on the vast majority of

afforestation sites in the LMRAV (Haines 1995, Kennedy

1990, Stanturf and others 1998). Although a multitude of

cost-share programs are available to private landowners,

most incentives target public environmental goals such as

conservation of highly erodible or otherwise delicate soil

(Haines 1995, Kennedy 1990). Even when program

objectives allow for “creation of wildlife habitat,” “promotion of

biodiversity,” or “production of sustainable timber harvest,”

these objectives often offer insufficient focus to ensure

optimal management of future stands for specific outputs

(Wilson and others, in press). The landowner should clearly

define singular or multiple objectives; for example, a

management objective for carbon sequestration would

describe the desired accumulated carbon or biomass output

in tons per acre per year over the specified rotation. Without

explicit objectives, the forester cannot properly determine

planting density, species assignments, site preparation

requirements, or postplanting cultural practices; nor can he

or she evaluate success of the afforestation effort. The

failure to specify management objectives has led to

wholesale establishment of stands without clear description

of future management pathways. Our observation suggests

that some landowners may be making uninformed decisions.

The current approach does not address sustainability issues

and could prove costly if the future forest structure is not

compatible with desired outputs.

Species–Site Selections

Alluvial floodplain forests exhibit high species richness and

spatial diversity of vegetational communities (Kellison and

others 1998, Meadows and Nowacki 1996). Bottomland

hardwood forests are comprised of more than 70 endemic

tree species along with numerous vines, shrubs, and

herbaceous species (Carter 1978, Putnam and others 1960,

Tanner 1986). A wide array of edaphic and hydrologic

conditions sculpted by the erosional and depositional

processes of rivers provide the foundation for vegetational

diversity in alluvial floodplains. Site types range from

permanently inundated sloughs with very poorly drained,

heavy clay soils to rarely inundated ridges of well-drained,

sandy loams (Stanturf and Schoenholtz 1998). Since the

early 1900s, studies have associated tree species with

various site types (Meadows and Nowacki 1996, Putnam

and others 1960, Tanner 1986). It follows that the suitability

of the species assigned to a given site will largely determine

initial and long-term afforestation success, the trajectory of

stand development, site productivity, future management

opportunities, and costs. Some of these relationships appear

in reports by Baker (1977), Dicke and Toliver (1987), Krinard

and Johnson (1985), Stine and others (1995), and Williams

and others (1993).

Afforestation information is available to assist the forester

with species-site prescriptions. Useful sources include “A

Practical Field Method of Site Evaluation for Commercially

Important Southern Hardwoods” by Baker and Broadfoot

(1979), “Hardwood Suitability for and Properties of Important

Midsouth Soils” by Broadfoot (1976), and county soil series

manuals published by the Natural Resource Conservation

Service. In addition, basic information characterizing

physical and chemical soil properties can identify soil texture

and drainage classes, plow pan development, nutrient

deficiencies or factors such as pH that regulate nutrient

uptake. Surveys of adjacent forested stands to determine

local abundance of desirable species can inform decisions

on species assignments (Groninger and others 1999). In

practice, though, market availability of planting stock is

probably the most prevalent factor-driving species

assignments on afforestation sites. Fewer than 25 of the 70-

plus native bottomland hardwood species are available

through established commercial nurseries on a yearly basis

[(Personal communication. 2000. Sam Campbell, Nursery

Manager, Molpus Timberlands Mgmt. L.L.C., 29650

Comstock, Elberta, AL 36530); (Personal communication.

2000. David McCain, Nursery Manager, Delta View Nursery,

Route 1 Box 28, Old Highway 61 South, Leland, MS 38756);

(Personal communication. 2000. Randy Rentz, Nursery

Manager, Columbia Nursery, P.O. Box 647, Columbia, LA

71418); (Personal communication. 2000. Gary Schaefer,

Nursery Manager, Winona Nursery, Route 3, Box 83,

Winona, MS 38967)]. However, some nursery managers will

raise custom seedlings of other species if contracted.

Site Preparation

Site preparation can be vital to afforestation of former

agricultural land. Treatments can condition the seed or

seedling bed; decrease competing or undesirable

vegetation, such as exotic pests; reduce herbivore habitat;

improve nutrient availability; and improve access on the site

for the planting operation (Baker and Blackmon 1978;

Kennedy 1981a, 1993). Benefits are typically realized

though increased survival and improved early growth of

hardwood planting stock (Baker and Blackmon 1978, Ezell

and Catchot 1998, Russell and others 1998). The wide array

of conditions on former agricultural fields precludes

wholesale prescription of site preparation practices. Rather,

the landowner’s objectives and the condition of the field

determine the appropriate level of site preparation for a

given tract. For fields immediately out of crop production, site

preparation is generally not necessary unless objectives and

site conditions make it desirable to break up a hard pan or

compacted soil, broadcast a pre-emergent herbicide

application for weed control, or incorporate fertilizer into the

planting site. Fertilization, for example, can consistently

boost growth of hardwood reproduction on former

agricultural sites, because long-term agricultural production

significantly depletes soil organic matter and associated

nutrients (Francis 1985, Houston and Buckner 1989). Such

practices are common if fiber production, timber production,

or biomass production are identified as primary

management objectives (Joslin and Schoenholtz 1998,

Kennedy 1981a, Thornton and others 1998, Yeiser 1999),

but also have merit where other objectives target early stand

growth and development.

Depending on the length of the uncultivated period and the

rate of succession, fields removed from cultivation for more

than a year prior to planting will present a range of

herbaceous and woody vine, shrub, or tree competition. It

may be desirable to control advance vegetation prior to

planting, and site preparation practices for such fields can

be accomplished with mechanical or chemical methods. In

the LMRAV, multiple-pass disking has been effective to bust

dense sod, improve soil aeration, and promote water
77

infiltration (Baker and Blackman 1978, Kennedy 1990,

McKnight 1970). Following years of cultivation, subsoil or

deep plowing to 16 to 20 in. is effective in breaking plow

pans that may develop. This practice, which is generally

necessary for establishment of fast-growing species, such

as eastern cottonwood (Populus deltoids Bartr. Ex Marsh)

(McKnight 1970), improves aeration and allows the

regeneration to exploit a greater soil volume.

Chemical site preparation and new dormant season weed

control applications show promise for relatively inexpensive

early control of herbaceous weeds for improved survival and

growth in hardwood plantations (Ezell 1995, 1999; Ezell and

Catchot 1998; Ezell and others 1999). Chemical methods of

site preparation offer the forester an ability to apply weed

control during periods when site conditions prevent use of

mechanical practices. Prior to herbicide application, mowing

or burning the field and allowing for a uniform regrowth of

vegetation can improve efficacy (Miller 1993).

In the LMRAV, mowing is a common site preparation

technique on afforestation projects sponsored by

governmental cost-share programs. This practice improves

planter access on afforestation sites that have not received

cultivation for several years, but mowing probably does little

to reduce weed competition or herbivory (Houston and

Buckner 1989). In fact, empirical studies rarely demonstrate

improved survival or growth of hardwood regeneration

following mowing for site preparation and/or subsequent

weed control (Houston and Buckner 1989, Kennedy 1981b,

Schweitzer and others 1999). Kennedy (1981b) reasoned

that mowing is not effective for improving growth or survival

because it does not reduce competition for soil water or

nutrients. Prescribed burning, a more economical practice

than mowing, can also be used to improve planter access on

afforestation sites. However, the use of prescribed fire

requires training, and liability related to smoke management

may limit the use of burning in some regions.

In the LMRAV, it is common to omit site preparation on some

sites to accommodate use of heavy equipment in machine

planting. To reduce site damage on saturated soil, competing

vegetation and the increased risk of herbivory are accepted

in a tradeoff with improved trafficability and planting machine

function.

Planting Stock Types, Size, and Procurement

Planting stocks used for afforestation in the LMRAV include

seed, bare-root seedlings, containerized seedlings, and

cuttings. Hard mast species, which can be successfully

established with seed on an operational scale include

several of the red oak species (Quercus nigra Linnaeus, Q.

phellos Linnaeus, Q. shumardii Buckley, Q. pagoda

Rafinesque, Q. nuttallii Palmer); white oak species (Q. lyrata

Walter, Q. michauxii Nuttall); common persimmon

(Diospyros virginiana Linnaeus); and sweet pecan [Carya

illinoensis (Wang.) K. Koch] (Johnson and Krinard 1985,

Stanturf and others 1998). Bottomland species that can be

established vegetatively with cuttings include eastern

cottonwood, American sycamore (Platanus occidentalis

Linnaeus), green ash (Fraxinus pennsylvanica Marshall),

and black willow (Salix nigra Marshall) (Kennedy 1977,

McKnight 1970, Stanturf and others 1998). The species

listed above and others not listed can be established as

bare-root or containerized seedlings. Choice of stock type

should be determined by management objectives, site

preparation practices, species decisions, planting window,

and market availability. King and Keeland (1999) estimated

that over 67 percent of the public land and cost-share

plantings in the LMRAV have been established with 1-0,

bare-root seedlings.

Size and quality of bare-root planting stock can be of major

importance in determining establishment success and early

growth of tree seedlings (Land 1983, Thompson and Schultz

1995). Because of differing growth rates, growth habits, i.e.

indeterminate, semi-determinate, or determinate, and

biomass accumulation patterns (Dickson 1994; Hodges and

Gardiner 1993; Long and Jones 1993, 1996), bottomland

hardwoods exhibit a wide range of interspecific seedling

morphologies. Early researchers working on bottomland

hardwood regeneration identified desirable seedlings as

having a shoot length of 30 to 36 in. and a root-collar

diameter of 1/4 to 3/8 in. or larger (Kennedy 1981a,

McKnight and Johnson 1980). However, definitive guidelines

defining optimal seedling dimensions for bottomland

hardwood species, particularly concerning factors that

reduce survival and growth such as competing vegetation

and flooding, have not been developed or published.

Research on quality seedling production in nursery beds has

revealed that certain practices can improve outplanting

performance, especially on harsh sites. In his review of

existing literature, South (1998) concluded that proper top

pruning of hardwood seedlings can significantly boost

outplanting survival (range 3 to 42 percent). Top pruning may

benefit the seedling by improving its root-weight ratio, while

it also helps the planter because top-pruned seedlings are

easier to handle. In addition to potential gains in survival,

stimulated height growth of seedlings from moderate top

pruning quickly compensates for the lost height from pruning

(Adams 1985, Meadows and Toliver 1987). Moderate root

pruning can also facilitate planting without significantly

reducing survival or growth (Toliver and others 1980). Yet,

root pruning should be approached cautiously because

excessive pruning will negatively alter root-weight ratio and

reduce carbohydrate reserves needed by the seedling to

survive lifting and transplanting. Kennedy (1993) suggested

that root systems of oak seedlings should be pruned to no

shorter than 8 in.

In addition to seedling size and handling practices,

morphological traits including the number of first-order

lateral roots can have a profound effect on early survival and

growth of hardwood seedlings (Thompson and Schultz

1995). Research conducted by Kormanic and colleagues

provides clear evidence linking seedling out-planting

performance and the inherited expression of first-order

lateral root proliferation (Kormanik 1986, 1989; Kormanik

and Ruehle 1987; Kormanik and others 1998; Thompson

and Schultz 1995). Based on their observations, Clark and

others (2000), Kormanik and Ruehle (1987), and Johnson

(1984) suggested that fewer than 40 percent of oak

seedlings lifted from nursery beds are suitable planting stock

based on lateral root development. Although first-order

lateral roots are strongly controlled by genetics, research
78

has demonstrated that their development can be increased

by growing seedlings at relatively low nursery bed densities

(Dey and Buchanan 1995). Most planting operations in the

LMRAV do not consider seedling morphology. Operational

programs generally target a shoot length of 18 to 24 in. and

a root-collar diameter of 3/8 in. as the minimal seedling size.

Clearly, managers need empirical data defining optimal

seedling dimensions and morphological traits to support

efficient planting of a diverse array of bottomland hardwood

species.

Few studies have examined the transfer of seed within the

southern hardwood region, but available evidence has

revealed provenance and family-within-provenance

differences for survival and growth of common species,

including cherrybark oak (Q. falcata var. pagodifolia Ell.),

American sycamore, and eastern cottonwood (Greene and

others 1991, Jokela and Mohn 1976, Land 1983). These

studies suggest that survival and growth can be increased

through provenance selections, but they also illustrate the

hazards of indiscriminate seed transfer. For example, Dicke

and Toliver (1987) observed a 30-percent range in survival

within cherrybark oak families at age 5. Transfer of seed to

different regions may be a concern, as well as establishing

upland ecotypes on bottomland soils. For example, Keeley

(1979) demonstrated that blackgum (Nyssa sylvatica

Marsh.) ecotypes selected along a flooding gradient

exhibited differing physiology, biomass accumulation

patterns, and survival rates. On the other hand, short-term

data presented by Yuceer and others (1998) revealed no

distinct differences in survival or growth of upland versus

bottomland sources of cherrybark oak. In practice, few

foresters in the LMRAV specify seed source constraints in

purchasing agreements. This lack of quality control or use of

certified seed in afforestation projects could potentially

reduce establishment success, productivity, and forest

health. Ideally, foresters should avoid transfer of seed

collected from other regions and site types until adequate

protocols for seed transfer are established. Morgenstern

(1996) provides conceptual details for establishing seed

transfer protocols for forest tree species. Interestingly, most

other developed countries and large companies have in

place transfer protocols and seed-certification programs for

the forests they manage.

Seed and Seedling Storage

Because seedlings are the predominant stock type currently

used in afforestation projects of the LMRAV, the remainder

of this manuscript will concentrate on practices and

techniques appropriate for them. Bonner and Vozzo (1987),

Bonner and others (1994), and Schopmeyer (1974) have

thoroughly discussed suitable techniques for collecting and

storing seed of bottomland hardwood species. Allen and

Kennedy (1989) and Stanturf and others (1998) describe

direct seeding techniques.

Bare-root seedlings should be lifted when dormant and

directly transferred to storage under refrigeration at 34 to 38

o

F. To maintain seedling viability in cold storage, seedling

bags should receive ample ventilation and moisture-content

monitoring. Mobile cold-storage facilities are readily available

for lease and most large-scale contractors maintain on-site

cold storage facilities during active planting. Such practices

enable operators to maintain seedling dormancy and

viability until time of planting.

Planting Seedlings

To establish reproduction on afforestation sites, contractors

operating in the LMRAV utilize crews of hand and machine

planters. Few studies have directly compared establishment

success rates between hand planting and machine planting

(Russell and others 1998). However, the authors have

observed that either method can be sufficiently effective if

experienced, conscientious personnel oversee the planting

job. We discuss techniques, advantages, and disadvantages

of each method below.

Hand planting—Hand-planting techniques originally

employed to establish large-scale hardwood plantations

were generally borrowed from conifer plantations. These

practices were generally not applicable to hardwood

plantation establishment in the LMRAV because of the

relatively large root system characteristic of most hardwood

seedlings and the often saturated, heavy clay alluvial soils.

Hardwood seedlings should be planted such that the

apparent root collar is at least 1 to 2 in. below the soil

surface. This practice helps ensure that all lateral roots are

sufficiently covered, and it can improve the sprouting

potential of seedlings, primarily oaks, that herbivores have

clipped. Hardwood seedlings typically have the taproot

pruned to about 8 in. and the laterals to 4 to 6 in. A planting

tool is needed with a blade at least 10- to 12-in. long by 6- to

8-in. wide. The type of dibble or planting shovel varies

among contractors, and often the same tool will not work

well on all sites due to soil and/or moisture conditions.

Because of the time and care required to plant large

seedlings properly in saturated soil, some contractors pay

their planters by the hour rather than by the number of

seedlings planted. A hand-planting crew of 20 people can

usually plant over 130 ac per day (about 2,000 seedlings per

planter at 300 seedlings per acre). This rate is quicker than a

machine-planting crew with one tractor. However, because

hand planting is labor intensive and requires more

administrative supervision and logistical planning to keep

planting crews active, it can be more expensive than

machine planting.

Machine planting—Machine planters for hardwoods are

largely similar to conifer planters, with modified packing

wheels and coulters to allow for planting of larger seedlings.

In addition, most operators modify stock planters to

accommodate their specific planting needs. Planting

machines are normally pulled by four-wheel drive, rubbertired tractors with a minimum rating of 175 horsepower. If soil

conditions are favorable, machine planting can be more

consistent for large seedlings with well-developed root

systems, and machine planting is generally not as expensive

as hand planting based on cost per seedling. A singlemachine planter crew can plant about 15 to 20 ac per day if

soil conditions are ideal. However, water saturated, heavy

clay soil typical of some alluvial floodplain sites can hamper

progress of machine planting, and the heavy equipment

required for machine planting can damage afforestation sites

by creating ruts. Furthermore, if soil conditions are not ideal,

the slit created by the planting machine is often never closed

near the lower reaches of the foot or coulter blade and may
79

serve as a site of cracking under dry conditions in the

smectitic soils of the LMRAV. Machine planting also

increases the minimal distance between rows, and may

damage growing stock if seedlings are being planted

supplemental to partial failures or volunteer regeneration.

Planting Job Inspections

Ongoing inspection is necessary to ensure proper seedling

handling, planting, and spacing. Viability problems from

improper handling or storage are near impossible to detect

after planting. Inspections also allow for real-time correction

of planting and spacing mistakes. Walk-through inspections

enable the forester to verify seedling condition and

appropriate root pruning. Establishing fixed-radius plots

behind the planting crew, the forester can monitor planting

density, seedling size, planting depth, and general quality.

Some choose to routinely sample one 0.02-ac plot for every

10 ac planted. However, seedling spacing should be

considered when determining the size of fixed-radius plots,

while sampling intensity will depend on the project area, site

heterogeneity, and the consistency of the planting crew.

Postplanting Cultural and Protection Practices

Postplanting cultural and protection practices improve

seedling survival, early growth, and plantation integrity.

Postplanting cultural practices primarily target competition

control as a means of boosting survival and improving

seedling growth, but irrigation and fertilization practices may

increase as future demands for hardwood fiber increase

(Francis 1985; Houston and Buckner 1989; Kennedy 1981a,

1981b, 1993; Schweitzer and others 1999; Yeiser 1999). In

spite of the demonstrated biological benefits, cost-benefit

analyses of postplanting operations have not been

conducted to project their financial benefits. However, the

additional costs of the practices may be justified if they

prevent plantation failure during drought or herbivore

damage, or if they significantly decrease rotation length as in

the case of disking operations in short-rotation woody crops.

In practice, few afforestation foresters prescribe postplanting

cultural treatments unless fiber or timber production is a

primary management objective.

Competition control in hardwood plantations can be

accomplished with mechanical or chemical methods, or with

using mulch material. Mechanical methods of competition

control primarily include mowing and disking. Because

mowing does not reduce belowground competition for soil

water and nutrients, hardwood reproduction generally does

not respond (Houston and Buckner 1989, Kennedy 1981b,

Schweitzer and others 1999). Mowing may only be practical

where the forester wishes to slow down development of

invasive woody species.

Although both share similar costs, disking is generally more

effective than mowing for controlling competing vegetation.

Several bottomland species including sweet pecan, Nuttall

oak, green ash, American sycamore, eastern cottonwood,

and sweetgum (Liquidambar styraciflua L.) respond

favorably to disking (Houston and Buckner 1989, Kennedy

1981b, Schweitzer and Stanturf 1999). In addition to

increasing aeration and moisture infiltration into soil, disking

improves nutrient status and subsequently growth of

hardwood reproduction (Kennedy 1981b). Gains in survival

and growth derived from disking often come from early stand

development, e.g., quicker advancement to canopy closure

and self-pruning. However, excessive disking or disking too

deep can prune roots excessively and reduce tree growth

(Schweitzer and Stanturf 1999).

Recent research on plantation establishment has identified

several herbicide tank mixes suitable for use with bottomland

hardwood species (Ezell 1999, Ezell and Catchot 1998,

Ezell and others 1999, Russell and others 1998). Vegetation

control with herbicides can effectively increase growth of

bottomland hardwood seedlings (Miller 1993, Russell and

others 1998) and may provide the most cost-effective control

of competing vegetation in relatively large, hardwood

plantations. However, most tank mixes are best suited for

controlling grass and some broadleaf herbaceous species,

and chemical technology is not available for woody vines,

shrubs, or trees in established plantations. Chemical control

of undesirable woody species can only be attained with

directed applications of suitable herbicides with appropriate

measures taken to minimize spray drift and contact with crop

species (Leininger and McCasland 1998, Miller 1993). Sites

occupied by resilient vine species, such as ladies’-eardrops

(Brunnichia cirrhosa Banks), trumpet creeper [Campsis

radicans (L.) Seemann], and peppervine [Ampelopsis

arborea (L.) Koehne], may require 2 or more years of

treatment before afforestation.

Mulching is generally more expensive and more

cumbersome than other methods of vegetation control, but it

can provide long-term efficacy resulting in dramatic gains in

survival and growth during the initial stages of stand

development (Adams 1997, Windell and Haywood 1996).

Limited research has demonstrated promising gains in early

growth for mulched common persimmon, green ash, Nuttall

oak, cherrybark oak, and water oak (Adams 1997,

Schweitzer and others 1999). Various organic and synthetic

mulch materials are commercially available, but a manager

should consider ease of application, durability of the

material, maintenance requirements, effectiveness, and cost

(Haywood 1999, Windell and Haywood 1996). Mulch use

may increase on wetland sites not amenable to mechanical

or chemical control.

In addition to improving early survival and growth of

seedlings, control of herbaceous vegetation can reduce

herbivory by modifying herbivore use of old field habitats

(Paul B. Hamel. 1995. Files/Sharkey/mammals. On file with:

U.S. Department of Agriculture, Forest Service, Southern

Research Station, Southern Hardwoods Laboratory, P.O. Box

227, Stoneville, MS 38776). White-tailed deer (Odocoileus

virginianus Zimmerman), rodents (including Sigmodon

hispidus Say and Ord), rabbits (Sylvilagus spp.), beaver

(Castor canadensis Kuhl), and nutria (Myocastor coypus

Molina) can be primary damaging agents in bottomland

hardwood plantations (Burkett and Williams 1998, Conner

and Toliver 1990, Conner and others 1999, King and

Keeland 1999, McKnight 1970). Animal damage can range

from mild, with little effect on planted seedlings, to severe, in

which high densities of herbivores decimate young tree

plantations (Conner and Toliver 1990). Aside from

modification of habitat, which is effective on rodents,
80

seedling protection or herbivore eradication practices may

discourage herbivory.

Shelters can increase seedling survival where herbivory

limits establishment (Conner and others 1999, Graveline and

others 1998, Strange and Shea 1998). Several different

styles of seedling shelters are available commercially, and

selection of style and size will depend on the size of

seedlings, expected herbivory type, costs, and assembly

and installation requirements (Windell 1991). Some tree

shelters also provide a favorable microclimate for improved

early tree growth (Schweitzer and others 1999, Tuley 1985).

Shelters can facilitate growth by moderating the light

environment, reducing seedling transpiration rates,

increasing temperature, and increasing carbon dioxide

(Tuley 1985, Windell 1991). However, early gains in height

growth are often due to temporary shifts in biomass

accumulation and are not always maintained after seedlings

grow above the shelters (Clatterbuck 1999, Mullins and

others 1998). Besides their high costs, shelters are easily

knocked down or swept away by floodwaters. These

drawbacks limit the use of shelters to sites of severe

herbivory. Perpetual eradication practices may most

effectively curtail severe herbivory by beaver and nutria.

Other protection in established plantations involves control

of insect or disease pests, fire prevention and suppression,

and floodwater management. Insects and diseases can

reduce plantation health and can render planted stock

vulnerable to other stress. For example, young plantations

of eastern cottonwood cultured for rapid biomass production

may require control of several pests including the

cottonwood leaf beetle (Chrysomela scripta Fabricius) and

the cottonwood borer [Plectrodera scalator (Fabricius)]

(Solomon 1985). Preventative practices such as selection of

resistant seed sources or clones may reduce damage by

insects and disease (Cooper and others 1977, Kellison

1994, Nebeker and others 1985); direct cultural, chemical,

or microbial techniques may also eradicate pests (Solomon

1985, Solomon and others 1997). In several useful

handbooks, Solomon and his colleagues describe major

insect pests and diseases of common bottomland tree

species including cottonwood, green ash, sycamore, and

the oak species (Leininger and others 1999; Morris and

others 1975; Solomon 1995; Solomon and others 1993,

1997).

Wildfire can destroy young hardwood plantations and reduce

stem quality on stump spouts. As a precautionary measure

against wildfire, Kennedy (1993) suggested maintenance of

fire lanes around all plantations. If fire sweeps though a

hardwood plantation, a site inventory must determine the

extent of damage and the necessary management.

Although most bottomland hardwood species exhibit some

level of tolerance to anaerobic soil, long-term inundation

during the growing season can harm all but the most floodtolerant species (Baker 1977, Hook 1984). Monitoring and

control of floodwater depth and duration are necessary if

survival of young hardwood seedlings is at stake. Where

flooding is desirable for waterfowl habitat, floodwater

removal before the active growing season will usually reduce

stress on seedlings. Additionally, by increasing soil moisture

availability during the potentially dry summer months, wellmanaged impoundments may improve seedling survival or

growth (Broadfoot 1967).

Postplanting Survival and Growth Monitoring

Comparing seedling survival and growth to the A Priori

definition of success can determine success of the planting

effort. The landowner’s management objectives, the type of

plantation, e.g., pure versus mixed species, availability of

preexisting data, and the costs of acquiring new data will

help determine sampling interval, timing, and measurement

intervals (Curtis 1983). However, prior to postplanting

assessments, baseline information on plantation

establishment will be vitally important to the afforestation

forester. Information such as seed source, seedling size and

condition, seedling lifting, shipment and storage history, soil

and atmospheric conditions during planting, planting

methods, planting contractor, site preparation activities, and

planting date can identify the source of problems or

successes. Postplanting assessment and monitoring

techniques vary widely among landowners and public

agencies, but they may often include sample transects,

permanent sample plots, photodocumentation, and periodic

aerial photography.

PLANTATION DESIGN

Hardwood plantations on former agricultural fields in the

LMRAV range from single-species to mixed-species

plantings. The afforestation forester should select a

particular plantation type based on the desired outputs

defined by management objectives (Daniel and others

1979). Single-species plantations, or monocultures, are

often the most efficient plantation type for optimizing a single

output, e.g., fiber production or soil amelioration. Singlespecies stands allow efficient cultural practices, more

predictable stand-development patterns, and more

predictable yields (Smith 1986). In the LMRAV, the native

soft broadleaf species that exhibit indeterminate growth

patterns are well suited for single-species stands. Perhaps

eastern cottonwood plantations, cultivated for high-quality

printing fiber, are the most extensive single-species

plantations in the LMRAV (Krinard and Johnson 1980). In

recent years, scientists in other regions have demonstrated

the value of fast-growing, single-species plantations as

catalysts for rehabilitating degraded forest ecosystems

(Parrotta and others 1997). In this role, rapidly grown aboveand belowground biomass stabilize soil, increase soil

organic matter, nutrient or water-holding capacity, develop

an understory microclimate that promotes establishment of

native species, and develop habitat for native fauna (Fisher

1995; Lugo 1997; Mapa 1995; Parrotta 1992, 1999). Singlespecies plantations often do not produce high-quality

sawtimber because most valuable species such as the oaks

generally develop their highest vigor and quality in stands

providing interspecific competition (Lockhart and Hodges

1998). Some managers may assume that single-species

stands provide poor wildlife habitat, but homogeneous

stands of eastern cottonwood, black willow, sandbar willow

(S. exigua Nutt.), and baldcypress [Taxodium distichum (L.)

Rich] occurring naturally along the Mississippi River

contribute to landscape diversity and provide critical habitat

for various wildlife species.
81

Mixed-species plantations can include various arrangements

of multiple species in true mixtures or intercropping mixtures

(Goelz 1995a). Potential benefits of mixed-species stands

versus single-species stands can increase pest resistance,

productivity in a vertically stratified stand, product diversity,

crop tree quality, and canopy species diversity (Goelz

1995b, Smith 1986). True mixtures generally consist of

randomly or systematically assigned species combinations

established at the same time. Some mixed plantations are

established with species of similar growth rates and

developmental patterns (Goelz 1995a), but most successful

mixtures require species that will stratify within the forest

canopy (Smith 1986). Stressing these points for bottomland

hardwood plantations, Lockhart and Hodges (1998) cited

work on mixed-species stand development by Clatterbuck

and Hodges (1988) and Clatterbuck and others (1987).

Lockhart and others (1999) also indicated that stand

development processes in well-designed species mixtures

will be similar to developmental tracts observed in natural

patterns. Most current afforestation practices under

governmental cost-share programs attempt to establish true

species mixtures to provide stand-level species diversity.

Unfortunately, many plantations are established without

consideration for the developmental trajectories and

competitive interactions of individual species comprising the

mixed plantation (Lockhart and Hodges 1998).

Establishing species that exhibit very different growth rates

can create intercropping mixtures. Such mixtures may

provide different products such as a commercial timber

species intercropped with a nitrogen-fixing species (Goelz

1995a). In the LMRAV, scientists and land managers have

developed an intercropping scheme using the earlysuccessional eastern cottonwood as a nurse species for the

slower growing, disturbance-dependent Nuttall oak

(Schweitzer and others 1997, Twedt and Portwood 1997). Its

very fast early growth, sparse crown architecture, and its

suitability to intensive culture make eastern cottonwood a

viable candidate as a nurse species. Potential benefits of the

eastern cottonwood-Nuttall oak intercropping could include

rapid rehabilitation of soil quality, rapid development of

vertical structure for faunal habitat, early financial return on

the rehabilitation investment, and development of an

understory favorable for oak seedlings and other native

woody species. Intercropping systems show potential for

providing multiple ecological and landowner benefits in the

LMRAV. Future research scheduled by the lead author and

cooperators will examine development of other intercropping

systems to extend application on a variety of bottomland site

types, e.g., use of black willow as a nurse for other species

on hydric sites.

MANAGEMENT CONSIDERATIONS

The LMRAV is currently experiencing extensive afforestation

of former agricultural fields on sites that historically

supported bottomland hardwood forests. Projections indicate

that the current pace may be maintained through the next

decade, resulting in hundreds of thousands of acres in

bottomland hardwood plantations. In summarizing our review

of literature, techniques, and practices, it became apparent

that several fundamental components of afforestation were

generally lacking in most regeneration practices currently

performed in the LMRAV. Developing some of these missing

components will require additional research, but others will

require only an extension of current knowledge or

application of conservation principles. Four fundamentally

vital components should be more deeply incorporated into

21st century, state-of-the-art afforestation activities in the

LMRAV:

(1) definition of specific landowner management objectives,

(2) establishment of stock size and quality guidelines,

(3) development of protocols for transfer of genetic

material, and

(4) application of silvicultural and ecological principles in

plantation establishment.

Incorporating these basic components will enable

landowners, natural resource managers, and the general

public a method of evaluating success of these afforestation

activities and should improve afforestation efficiency,

ecosystem health, and resource sustainability in the LMRAV.

LITERATURE CITED

Adams, J.C. 1985. Severe top pruning improves water oak seedling

growth. In: Shoulders, Eugene, ed. Proceedings of the third

biennial southern silvicultural research conference; 1984

November 7–8; Atlanta. Gen. Tech. Rep. SO–54. New Orleans:

U.S. Department of Agriculture, Forest Service, Southern Forest

Experiment Station: 1–3.

Adams, J.C. 1997. Mulching improves early growth of four oak

species in plantation establishment. Southern Journal of Applied

Forestry. 21: 44–46.

Allen, J.A. 1997. Reforestation of bottomland hardwoods and the

issue of woody species diversity. Restoration Ecology. 5: 125–

134.

Allen, J.A.; Kennedy, H.E., Jr. 1989. Bottomland hardwood reforestation in the Lower Mississippi Valley. Slidell, LA: U.S. Department

of the Interior, Fish and Wildlife Service, National Wetlands

Research Center; Stoneville, MS: U.S. Department of Agriculture, Forest Service, Southern Forest Experiment Station. 28 p.

Baker, J.B. 1977. Tolerance of planted hardwoods to spring flooding.

Southern Journal of Applied Forestry. 1: 23–25.

Baker, J.B.; Blackmon, B.G. 1978. Summer fallowing - a simple

technique for improving old-field sites for cottonwood. Res. Pap.

SO–142. New Orleans: U.S. Department of Agriculture, Forest

Service, Southern Forest Experiment Station. 5 p.

Baker, J.B.; Broadfoot, W.M. 1979. A practical field method of site

evaluation for commercially important southern hardwoods. Gen.

Tech. Rep. SO–26. New Orleans: U.S. Department of Agriculture,

Forest Service, Southern Forest Experiment Station. 51 p.

Bonner, F.T.; Vozzo, J.A. 1987. Seed biology and technology of

Quercus. Gen. Tech. Rep. SO–66. New Orleans: U.S. Department of Agriculture, Forest Service, Southern Forest Experiment

Station. 21 p.

Bonner, F.T.; Vozzo, J.A.; Elam, W.W.; Land, S.B., Jr. 1994. Tree

seed technology training course. Gen. Tech. Rep. SO–107. New

Orleans: U.S. Department of Agriculture, Forest Service,

Southern Forest Experiment Station. 81 p.

Broadfoot, W.M. 1967. Shallow-water impoundments increases soil

moisture and growth of hardwoods. Soil Science Society of

America Proceedings. 31: 562–564.
82

Broadfoot, W.M. 1976. Hardwood suitability for and properties of

important midsouth soils. Res. Pap. SO–127. New Orleans: U.S.

Department of Agriculture, Forest Service, Southern Forest

Experiment Station. 84 p.

Burkett, V.; Williams, H. 1998. Effects of flooding regime, mycorrhizal inoculation and seedling treatment type on first-year

survival of Nuttall oak (Quercus nuttallii Palmer). In: Waldrop,

Thomas A., ed. Proceedings of the ninth biennial southern

silvicultural research conference; 1997 February 25–27;

Clemson, SC. Gen. Tech. Rep. SRS–20. Asheville, NC: U.S.

Department of Agriculture, Forest Service, Southern Research

Station: 289–294.

Carter, J.R., Jr. 1978. A floristic study of the Delta National Forest

and adjacent areas. Mississippi State, MS: Mississippi State

University. 79 p. M.S. thesis.

Clark, S.L.; Schlarbaum, S.E.; Kormanik, P.P. 2000. Visual grading

and quality of 1-0 northern red oak seedlings. Southern Journal

of Applied Forestry. 24: 93–97.

Clatterbuck, W.K. 1999. Effects of tree shelters on growth of

hardwood seedlings after seven growing seasons. In: Haywood,

James D., ed. Proceedings of the tenth biennial southern

silvicultural research conference; 1999 February 16–18;

Shreveport, LA. Gen. Tech. Rep. SRS–30. Asheville, NC: U.S.

Department of Agriculture, Forest Service, Southern Research

Station: 43–46.

Clatterbuck, W.K.; Hodges, J.D. 1988. Development of cherrybark

oak and sweet gum in mixed, even-aged bottomland stands in

central Mississippi, U.S.A. Canadian Journal of Forest Research.

18: 12–18.

Clatterbuck, W.K.; Oliver, C.D.; Burkhardt, E.C. 1987. The silvicultural potential of mixed stands of cherrybark oak and American

sycamore: spacing is the key. Southern Journal of Applied

Forestry. 11: 158–161.

Conner, W.H.; McLeod, K.W.; Inabinette, L.W. [and others]. 1999.

Successful planting of tree seedlings in wet areas. In: Haywood,

James D., ed. Proceedings of the tenth biennial southern

silvicultural research conference; 1999 February 16–18;

Shreveport, LA. Gen. Tech. Rep. SRS–30. Asheville, NC: U.S.

Department of Agriculture, Forest Service, Southern Research

Station: 201–204.

Conner, W.H.; Toliver, J.R. 1990. Observations on the regeneration

of baldcypress (Taxodium distichum [L.] Rich.) in Louisiana

swamps. Southern Journal of Applied Forestry. 14: 115–118.

Cooper, D.T.; Filer, T.H., Jr.; Wells, O.O. 1977. Geographic variation

in disease susceptibility of American sycamore. Southern

Journal of Applied Forestry. 1: 21–24.

Curtis, R.O. 1983. Procedures for establishment and maintaining

permanent plots for silvicultural and yield research. Gen. Tech.

Rep. PNW–155. Portland, OR: U.S. Department of Agriculture,

Forest Service, Pacific Northwest Forest and Range Experiment

Station. 56 p.

Daniel, T.W.; Helms, J.A.; Baker, F.S. 1979. Principles of silviculture.

2d ed. St. Louis: McGraw-Hill, Inc. 500 p.

Dey, D.; Buchanan, M. 1995. Red oak (Quercus rubra L.) acorn

collection, nursery culture and direct seeding: a literature review.

For. Res. Inf. Pap. 122. Sault Ste. Marie, Ontario: Ontario

Ministry of Natural Resources, Ontario Forest Research

Institute. 46 p.

Dicke, S.G.; Toliver, J.R. 1987. Response of cherrybark oak families

to different soil-site conditions. In: Phillips, Douglas R., comp.

Proceedings of the fourth biennial southern silvicultural research

conference; 1986 November 4–6; Atlanta. Gen. Tech. Rep. SE–

42. Asheville, NC: U.S. Department of Agriculture, Forest

Service, Southeastern Forest Experiment Station: 260–263.

Dickson, R.E. 1994. Height growth and episodic flushing in northern

red oak. In: Isebrands, J.G.; Dickson, R.E., comps. Biology and

silviculture of northern red oak in the north central region: a

synopsis. Gen. Tech. Rep. NC–173. St. Paul, MN: U.S. Department of Agriculture, Forest Service, North Central Forest

Experiment Station: 1–9.

Ezell, A.W. 1995. Importance of early season competition control in

establishing eastern cottonwood (Populus deltoides) plantations.

In: Edwards, M. Boyd, comp. Proceedings of the eighth biennial

southern silvicultural research conference; 1994 November 1–3;

Auburn, AL. Gen. Tech. Rep. SRS–1. Asheville, NC: U.S.

Department of Agriculture, Forest Service, Southern Research

Station: 94–97.

Ezell, A.W. 1999. Crop tolerance of Nuttall, water, and willow oaks to

preemergent applications of DPX–6447: second year results. In:

Haywood, James D., ed. Proceedings of the tenth biennial

southern silvicultural research conference; 1999 February 16–

18; Shreveport, LA. Gen. Tech. Rep. SRS–30. Asheville, NC: U.S.

Department of Agriculture, Forest Service, Southern Research

Station: 142–144.

Ezell, A.W.; Catchot, A.L., Jr. 1998. Competition control for

hardwood plantation establishment. In: Waldrop, Thomas A., ed.

Proceedings of the ninth biennial southern silvicultural research

conference; 1997 February 25–27; Clemson, SC. Gen. Tech.

Rep. SRS–20. Asheville, NC: U.S. Department of Agriculture,

Forest Service, Southern Research Station: 42–43.

Ezell, A.W.; Portwood, J.; Quicke, H. 1999. Pre- and postemergent

applications of imazaquin for herbaceous weed control in

eastern cottonwood plantations: second year results. In:

Haywood, James D., ed. Proceedings of the tenth biennial

southern silvicultural research conference; 1999 February 16–

18; Shreveport, LA. Gen. Tech. Rep. SRS–30. Asheville, NC: U.S.

Department of Agriculture, Forest Service, Southern Research

Station: 73–75.

Fisher, R.F. 1995. Amelioration of degraded rain forest soils by

plantations of native trees. Soil Science Society of America

Journal. 59: 544–549.

Francis, J.K. 1985. Bottomland hardwood fertilization – the

Stoneville experience. In: Shoulders, Eugene, ed. Proceedings of

the third biennial southern silvicultural research conference;

1984 November 7–8; Atlanta. Gen. Tech. Rep. SO–54. New

Orleans: U.S. Department of Agriculture, Forest Service,

Southern Forest Experiment Station: 346–350.

Goelz, J.C.G. 1995a. Establishment of mixed-species plantations of

bottomland hardwoods. In: Landin, M.C., ed. National interagency workshop on wetlands: technology advances for

wetlands science; 1995 April 3–7; New Orleans. Vicksburg, MS:

U.S. Army Corps of Engineers, Waterways Experiment Station:

178–180.

Goelz, J.C.G. 1995b. Experimental designs for mixed-species

plantations. In: Edwards, M. Boyd, comp. Proceedings of the

eighth biennial southern silvicultural research conference; 1994

November 1–3; Auburn, AL. Gen. Tech. Rep. SRS–1. Asheville,

NC: U.S. Department of Agriculture, Forest Service, Southern

Research Station: 559–563.
83

Graveline, B.D.; Wells, G.R.; Schlarbaum, S.; Fribourg, H.A. 1998.

Growth and protection of selected northern red oak seedlings

planted on old field sites. In: Waldrop, Thomas A., ed. Proceedings of the ninth biennial southern silvicultural research

conference; 1997 February 25–27; Clemson, SC. Gen. Tech.

Rep. SRS–20. Asheville, NC: U.S. Department of Agriculture,

Forest Service, Southern Research Station: 257–262.

Greene, T.A.; Lowe, W.J.; Stine, M. 1991. Volume production of six

cherrybark oak provenances in the western gulf region. In:

Coleman, Sandra S.; Neary, Daniel G., comps., eds. Proceedings of the sixth biennial southern silvicultural research

conference; 1990 October 30–November 1; Memphis, TN. Gen.

Tech. Rep. SE–70. Asheville, NC: U.S. Department of Agriculture, Forest Service, Southeastern Forest Experiment Station:

736–743.

Groninger, J.W.; Aust, W.M.; Miwa, M.; Stanturf, J.A. 1999. Tree

species-soil relationships on marginal soybean lands in the

Mississippi Delta. In: Haywood, James D., ed. Proceedings of the

tenth biennial southern silvicultural research conference; 1999

February 16–18; Shreveport, LA. Gen. Tech. Rep. SRS–30.

Asheville, NC: U.S. Department of Agriculture, Forest Service,

Southern Research Station: 205–209.

Haines, T. 1995. Federal and State forestry cost-share assistance

programs: structure, accomplishments, and future outlook. Res.

Pap. SO–295. New Orleans: U.S. Department of Agriculture,

Forest Service, Southern Forest Experiment Station. 18 p.

Haywood, J.D. 1999. Durability of selected mulches, their ability to

control weeds, and influence growth of loblolly pine seedlings.

New Forests. 18: 263–276.

Hodges, J.D.; Gardiner, E.S. 1993. Ecology and physiology of oak

regeneration. In: Loftis, David L.; McGee, Charles E., eds. Oak

regeneration: serious problems, practical recommendations:

Symposium proceedings; 1992 September 8–10; Knoxville, TN.

Gen. Tech. Rep. SE–84. Asheville, NC: U.S. Department of

Agriculture, Forest Service, Southeastern Forest Experiment

Station: 54–65.

Hook, D.D. 1984. Waterlogging tolerance of lowland tree species of

the South. Southern Journal of Applied Forestry. 8: 136–149.

Houston, A.E.; Buckner, E.R. 1989. Cultural treatments for the

establishment of bottomland hardwoods in west Tennessee. In:

Miller, James H., comp. Proceedings of the fifth biennial southern

silvicultural research conference; 1988 November 1–3; Gen.

Tech. Rep. SO–74. New Orleans: U.S. Department of Agriculture,

Forest Service, Southern Forest Experiment Station: 285–288.

Johnson, P.S. 1984. Responses of planted northern red oak to three

overstory treatments. Canadian Journal of Forest Research. 14:

536–542.

Johnson, R.L.; Krinard, R.M. 1985. Oak seeding on an adverse field

site. Res. Note SO–319. New Orleans: U.S. Department of

Agriculture, Forest Service, Southern Forest Experiment Station.

4 p.

Jokela, J.J.; Mohn, C.A. 1976. Geographic variation in eastern

cottonwood. In: Thielges, B.A.; Land, S. B., eds. Proceedings of

the symposium on eastern cottonwood and related species;

1976 September 28–October 2; St. Paul, MN. Sci. J. Ser. Pap.

1638. St. Paul, MN: University of Minnesota Agricultural

Experimental Station: 109–125.

Joslin, J.D.; Schoenholtz, S.H. 1998. Measuring the environmental

effects of converting cropland to short-rotation woody crops: a

research approach. Biomass and Bioenergy. 15: 301–311.

Keeley, J.E. 1979. Population differentiation along a flood frequency

gradient: physiological adaptations to flooding in Nyssa

sylvatica. Ecological Monographs. 49: 89–108.

Kellison, R.C. 1994. Genetic improvement of southern hardwoods -

1994 update. In: Landis, T.D.; Dumroese, R.K., tech. coords.

National proceedings, forest and conservation nursery associations. Gen. Tech. Rep. RM–257. Fort Collins, CO: U.S. Department of Agriculture, Forest Service, Rocky Mountain Forest and

Range Experiment Station: 20–25.

Kellison, R.C.; Young, M.J.; Braham, R.R.; Jones, E.J. 1998. Major

alluvial floodplains. In: Messina, M.G.; Conner, W.H., eds.

Southern forested wetlands - ecology and management. Boca

Raton, FL: CRC Press LLC: 291–341.

Kennedy, H.E., Jr. 1977. Planting depth and source affect survival of

planted green ash cuttings. Res. Note SO–224. New Orleans:

U.S. Department of Agriculture, Forest Service, Southern Forest

Experiment Station. 3 p.

Kennedy, H.E., Jr. 1981a. Bottomland hardwoods research on site

preparation, plantation establishment, and cultural treatments, at

the Southern Hardwoods Laboratory. In: Barnett, James P., ed.

Proceedings of the first biennial southern silvicultural research

conference; 1980 November 6–7; Atlanta. Gen. Tech. Rep. SO–

34. New Orleans: U.S. Department of Agriculture, Forest

Service, Southern Forest Experiment Station: 75–78.

Kennedy, H.E., Jr. 1981b. Foliar nutrient concentrations and

hardwood growth influenced by cultural treatments. Plant and

Soil. 63: 307–316.

Kennedy, H.E., Jr. 1990. Hardwood reforestation in the South:

landowners can benefit from Conservation Reserve Program

incentives. Res. Note SO–364. New Orleans: U.S. Department of

Agriculture, Forest Service, Southern Forest Experiment Station.

6 p.

Kennedy, H.E., Jr. 1993. Artificial regeneration of bottomland

hardwoods. In: Loftis, David L.; McGee, Charles E., eds. Oak

regeneration: serious problems, practical recommendations.

Symposium proceedings; 1992 September 8–10; Knoxville, TN.

Gen. Tech. Rep. SE–84. Asheville, NC: U.S. Department of

Agriculture, Forest Service, Southeastern Forest Experiment

Station: 241–249.

King, S.L.; Keeland, B.D. 1999. Evaluation of reforestation in the

Lower Mississippi River Alluvial Valley. Restoration Ecology. 7:

348–359.

Kormanik, P.P. 1986. Lateral root morphology as an expression of

sweetgum seedling quality. Forest Science. 32: 595–604.

Kormanik, P.P. 1989. Frequency distribution of first-order lateral

roots in forest tree seedlings: silvicultural implications. In: Miller,

James H., comp. Proceedings of the fifth biennial southern

silvicultural research conference; 1988 November 1–3; Gen.

Tech. Rep. SO–74. New Orleans: U.S. Department of Agriculture,

Forest Service, Southern Forest Experiment Station: 101–105.

Kormanik, P.P.; Ruehle, J.L. 1987. Lateral root development may

define nursery seedling quality. In: Phillips, Douglas R., comp.

Proceedings of the fourth biennial southern silvicultural research

conference; 1986 November 4–6; Atlanta. Gen. Tech. Rep. SE–

42. Asheville, NC: U.S. Department of Agriculture, Forest

Service, Southeastern Forest Experiment Station: 225–229.
84

Kormanik, P.P.; Sung, S.S.; Kass, D.J.; Schlarbaum, S. 1998. Effect

of seedling size and first-order-lateral roots on early development of northern red oak on mesic sites. In: Waldrop, Thomas

A., ed. Proceedings of the ninth biennial southern silvicultural

research conference; 1997 February 25–27; Clemson, SC. Gen.

Tech. Rep. SRS–20. Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station: 247–252.

Krinard, R.M.; Johnson, R.L. 1980. Fifteen years of cottonwood

plantation growth and yield. Southern Journal of Applied

Forestry. 4: 180–185.

Krinard, R.M.; Johnson, R.L. 1985. Eighteen-year development of

sweetgum (Liquidambar styraciflua L.) plantings on two sites.

Tree Planters’ Notes. 36: 6–8.

Land, S.B., Jr. 1983. Performance and G-E interactions of sycamore

established from cuttings and seedlings. In: Jones, Earle P., Jr.,

ed. Proceedings of the second biennial southern silvicultural

research conference; 1982 November 4–5; Atlanta. Gen. Tech.

Rep. SE–24. Asheville, NC: U.S. Department of Agriculture,

Forest Service, Southeastern Forest Experiment Station: 431–

440.

Leininger, T.D.; McCasland, C.S. 1998. Does paraquat cause stem

swellings in first-year cottonwood saplings? In: Waldrop, Thomas

A., ed. Proceedings of the ninth biennial southern silvicultural

research conference; 1997 February 25–27; Clemson, SC. Gen.

Tech. Rep. SRS–20. Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station: 44–48.

Leininger, T.D.; Solomon, J.D.; Wilson, A.D.; Schiff, N.M. 1999. A

guide to major insects, diseases, air pollution injury, and

chemical injury of sycamore. Gen. Tech. Rep. SRS–28. Asheville,

NC: U.S. Department of Agriculture, Forest Service, Southern

Research Station. 44 p.

Lockhart, B.R.; Ezell, A.W.; Hodges, J.D.; Clatterbuck, W.K. 1999.

Development of mixed cherrybark oak-sweetgum plantations

planted at different spacings in east-central Mississippi after 17

years. In: Haywood, James D., ed. Proceedings of the tenth

biennial southern silvicultural research conference; 1999

February 16–18; Shreveport, LA. Gen. Tech. Rep. SRS–30.

Asheville, NC: U.S. Department of Agriculture, Forest Service,

Southern Research Station: 103–106.

Lockhart, B.R.; Hodges, J.D. 1998. Bottomland red oak stand

development with implications for management and future

research. In: Waldrop, Thomas A., ed. Proceedings of the ninth

biennial southern silvicultural research conference; 1997

February 25–27; Clemson, SC. Gen. Tech. Rep. SRS–20.

Asheville, NC: U.S. Department of Agriculture, Forest Service,

Southern Research Station: 548–554.

Long, T.J.; Jones, R.H. 1993. Biomass partitioning in seedlings of

ten Quercus species native to the Southeastern United States.

In: Brissette, John C., ed. Proceedings of the seventh biennial

southern silvicultural research conference; 1992 November 17–

19; Mobile, AL. Gen. Tech. Rep. SO–93. New Orleans: U.S.

Department of Agriculture, Forest Service, Southern Forest

Experiment Station: 257–261.

Long, T.J.; Jones, R.H. 1996. Seedling growth strategies and seed

size effects in fourteen oak species native to different soil

moisture habitats. Trees. 11: 1–8.

Lugo, A.E. 1997. The apparent paradox of reestablishing species

richness on degraded lands with tree monocultures. Forest

Ecology and Management. 99: 9–19.

Mapa, R.B. 1995. Effect of reforestation using Tectona grandis on

infiltration and soil water retention. Forest Ecology and Management. 77: 119–125.

McKnight, J.S. 1970. Planting cottonwood cuttings for timber

production in the South. Res. Pap. SO–60. New Orleans: U.S.

Department of Agriculture, Forest Service, Southern Forest

Experiment Station. 17 p.

McKnight, J.S.; Johnson, R.L. 1980. Hardwood management in

southern bottomlands. Forest Farmer. 39: 30–39.

Meadows, J.S.; Nowacki, G.J. 1996. An old-growth definition for

eastern riverfront forests. Gen. Tech. Rep. SRS–4. Asheville, NC:

U.S. Department of Agriculture, Forest Service, Southern

Research Station. 7 p.

Meadows, J.S.; Toliver, J.R. 1987. Three-year response of pecan to

six methods of seedling establishment. Southern Journal of

Applied Forestry. 11: 56–59.

Miller, J.H. 1993. Oak plantation establishment using mechanical

burning, and herbicide treatments. In: Loftis, David L.; McGee,

Charles. E., eds. Oak regeneration: serious problems, practical

recommendations: Symposium proceedings; 1992 September 8–

10; Knoxville, TN. Gen. Tech. Rep. SE–84. Asheville, NC: U.S.

Department of Agriculture, Forest Service, Southeastern Forest

Experiment Station: 264–289.

Morgenstern, E.K. 1996. Geographic variation in forest trees:

genetic basis and application of knowledge in silviculture.

Vancouver, BC: UBC Press, University of British Columbia. 209

p.

Morris, R.C.; Filer, T.H.; Solomon, J.D. [and others]. 1975. Insects

and diseases of cottonwood. Gen. Tech. Rep. SO–8. New

Orleans: U.S. Department of Agriculture, Forest Service,

Southern Research Station. 37 p.

Mullins, J.A.; Buckner, E.R.; Hopper, G.; Percell, G.G. 1998. Early

growth of sheltered and unsheltered cherrybark oak established

by planting 1-0 bareroot and 1-0 containerized seedlings, and by

direct-seeding. In: Waldrop, Thomas A., ed. Proceedings of the

ninth biennial southern silvicultural research conference; 1997

February 25–27; Clemson, SC. Gen. Tech. Rep. SRS–20.

Asheville, NC: U.S. Department of Agriculture, Forest Service,

Southern Research Station: 234–237.

Nebeker, T.E.; Schmitt, J.J.; Solomon, J.D.; Honea, C.R. 1985.

Clonal resistance to and incidence of the poplar borer in

southern cottonwood plantations. In: Shoulders, Eugene, ed.

Proceedings of the third biennial southern silvicultural research

conference; 1984 November 7–8; Atlanta. Gen. Tech. Rep. SO–

54. New Orleans: U.S. Department of Agriculture, Forest

Service, Southern Forest Experiment Station: 247–251.

Parrotta, J.A. 1992. The role of plantation forests in rehabilitating

degraded tropical ecosystems. Agriculture, Ecosystems and

Environment. 41: 115–133.

Parrotta, J.A. 1999. Productivity, nutrient cycling, and succession in

single- and mixed-species plantations of Casuarina equisetifolia,

Eucalyptus robusta, and Leucaena leucocephala in Puerto Rico.

Forest Ecology and Management. 124: 45–77.

Parrotta, J.A.; Turnbull, J.W.; Jones, N. 1997. Catalyzing native

forest regeneration on degraded tropical lands. Forest Ecology

and Management. 99: 1–7.
85

Putnam, J.A.; Furnival, G.M.; McKnight, J.S. 1960. Management

and inventory of southern hardwoods. Agric. Handb. 181.

Washington, DC: U.S. Department of Agriculture, Forest Service.

102 p.

Russell, D.R., Jr.; Hodges, J.D.; Ezell, A.W. 1998. An evaluation of

hardwood reforestation methods on previously farmed lands in

central Alabama. In: Waldrop, Thomas A., ed. Proceedings of the

ninth biennial southern silvicultural research conference; 1997

February 25–27; Clemson, SC. Gen. Tech. Rep. SRS–20.

Asheville, NC: U.S. Department of Agriculture, Forest Service,

Southern Research Station: 272–276.

Schopmeyer, C.S. 1974. Seeds of woody plants in the United

States. Agric. Handb. 450. Washington, DC. U.S. Department of

Agriculture, Forest Service. 883 p.

Schweitzer, C.J.; Gardiner, E.S.; Stanturf, J.A.; Ezell, A.W. 1999.

Methods to improve establishment and growth of bottomland

hardwood artificial regeneration. In: Stringer, Jeffrey W.; Loftis,

David L., eds. Proceedings, 12th central hardwood forest

conference; 1999 February 28, March 1–2; Lexington, KY. Gen.

Tech. Rep. SRS–24. Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station: 209–214.

Schweitzer, C.J.; Stanturf, J.A. 1999. A comparison of large-scale

reforestation techniques commonly used on abandoned fields in

the Lower Mississippi Alluvial Valley. In: Haywood, James D., ed.

Proceedings of the tenth biennial southern silvicultural research

conference; 1999 February 16–18; Shreveport, LA. Gen. Tech.

Rep. SRS–30. Asheville, NC: U.S. Department of Agriculture,

Forest Service, Southern Research Station: 136–141.

Schweitzer, C.J.; Stanturf, J.A.; Shepard, J.P. [and others]. 1997.

Large-scale comparison of reforestation techniques commonly

used in the Lower Mississippi Alluvial Valley: first year results. In:

Pallardy, S.G; Cecich, R.A.; Garrett, H.G.; Johnson, P.S., eds.

Proceedings, 11th central hardwood forest conference; 1997

March 23–26; Columbia, MO. Gen. Tech. Rep. NC–188. St. Paul,

MN: U.S. Department of Agriculture, Forest Service, North

Central Forest Experiment Station: 313–320.

Solomon, J.D. 1985. Impact of insects on growth and development

of young cottonwood plantations. Res. Pap. SO–217. New

Orleans: U.S. Department of Agriculture, Forest Service,

Southern Forest Experiment Station. 6 p.

Solomon, J.D. 1995. Guide to insect borers of North American

broadleaf trees and shrubs. Agric. Handb. 706. Washington, DC:

U.S. Department of Agriculture, Forest Service. 735 p.

Solomon, J.D.; Leininger, T.D.; Wilson, A.D. [and others]. 1993. Ash

pests - a guide to major insects, diseases, air pollution injury,

and chemical injury. Gen. Tech. Rep. SO–96. New Orleans: U.S.

Department of Agriculture, Forest, Service, Southern Forest

Experiment Station. 45 p.

Solomon, J.D.; McCracken, F.I.; Anderson, R.L. [and others]. 1997.

Oak pests - a guide to major insects, diseases, air pollution and

chemical injury. Prot. Rep. R8–PR7. Atlanta: U.S. Department of

Agriculture, Forest Service, Southern Region. 69 p.

South, D.B. 1998. Effects of top-pruning on survival of southern

pines and hardwoods. In: Waldrop, Thomas A., ed. Proceedings

of the ninth biennial southern silvicultural research conference;

1997 February 25–27; Clemson, SC. Gen. Tech. Rep. SRS–20.

Asheville, NC: U.S. Department of Agriculture, Forest Service,

Southern Research Station: 3–8.

Smith, D.M. 1986. The practice of silviculture. 8th ed. New York: John

Wiley. 527 p.

Stanturf, J.A.; Schoenholtz, S.H. 1998. Soils and landforms. In:

Messina, M.G.; Conner, W.H., eds. Southern forested wetlands -

ecology and management. Boca Raton, FL: CRC Press LLC:

123–147.

Stanturf, J.A.; Schoenholtz, S.H.; Schweitzer, C.J.; Shepard, J.P.

2001. Achieving restoration success: myths in bottomland

hardwood forests. Restoration Ecology. 9(2): 189–200.

Stanturf, J.A.; Schweitzer, C.J.; Gardiner, E.S. 1998. Afforestation of

marginal agricultural land in the Lower Mississippi River Alluvial

Valley, U.S.A. Silva Fennica. 32: 281–297.

Stine, M.; Chambers, J.L.; Wilson, M.; Ribbeck, K. 1995. Twentyyear survival and growth of six bottomland hardwood species. In:

Edwards, M. Boyd, comp. Proceedings of the eighth biennial

southern silvicultural research conference; 1994 November 1–3;

Auburn, AL. Gen. Tech. Rep. SRS–1. Asheville, NC: U.S.

Department of Agriculture, Forest Service, Southern Research

Station: 500–502.

Strange, E.E.; Shea, K.L. 1998. Effects of deer browsing, fabric

mats, and tree shelters on Quercus rubra seedlings. Restoration

Ecology. 6: 29–34.

Tanner, J.T. 1986. Distribution of tree species in Louisiana bottomland forests. Castanea. 51: 168–174.

Thompson, J.R.; Schultz, R.C. 1995. Root system morphology of

Quercus rubra L. planting stock and 3-year field performance in

Iowa. New Forests. 9: 225–236.

Thornton, F.C.; Joslin, J.D.; Bock, B.R. [and others]. 1998. Environmental effects of growing woody crops on agricultural land: first

year effects on erosion, and water quality. Biomass and

Bioenergy. 15: 57–69.

Toliver, J.R.; Sparks, R.C.; Hansbrough, T. 1980. Effects of top and

lateral root pruning on survival and early growth - three bottomland hardwood tree species. Tree Planters’ Notes. 31: 13–15.

Tulley, G. 1985. The growth of young oak trees in shelters. Forestry.

58: 181–195.

Twedt, D.J.; Portwood, J. 1997. Bottomland hardwood reforestation

for neotropical migratory birds: are we missing the forest for the

trees? Wildlife Society Bulletin. 25: 647–652.

Williams, H.M.; Kleiss, B.A.; Humphrey, M.N.; Klimas, C.V. 1993.

First-year field performance of oak species with varying flood

tolerance planted on hydric and non-hydric soils. In: Brissette,

John C., ed. Proceedings of the seventh biennial southern

silvicultural research conference; 1992 November 17–19; Mobile,

AL. Gen. Tech. Rep. SO–93. New Orleans: U.S. Department of

Agriculture, Forest Service, Southern Forest Experiment Station:

409–414.

Wilson, R.R.; Oliver, J.M.; Twedt, D.J.; Uihlein, W.B., III. [In press].

Bottomland hardwood restoration in the Mississippi Alluvial

Valley: looking past the trees to see the forest. In: Fredrickson,

L., ed. Ecology and management of bottomland hardwood

systems symposium; 1999 March 11–13; Memphis, TN.

Columbia, MO: University of Missouri.

Windell, K. 1991. Tree shelters for seedling protection. Tech. Rep.

9124–2834– MTDC. Missoula, MT: U.S. Department of Agriculture, Forest Service, Technology and Development Program.

142 p.
86

Windell, K.; Haywood, J.D. 1996. Mulch mat materials for improved

tree establishment. Tech. Rep. 9624–2811–MTDC. Missoula, MT:

U.S. Department of Agriculture, Forest Service, Technology and

Development Center. 124 p.

Yeiser, J.L. 1999. Irrigated and unirrigated eastern cottonwood and

water oak in a short rotation fiber system on a former agricultural

site. In: Haywood, James D., ed. Proceedings of the tenth

biennial southern silvicultural research conference; 1999

February 16–18; Shreveport, LA. Gen. Tech. Rep. SRS–30.

Asheville, NC: U.S. Department of Agriculture, Forest Service,

Southern Research Station: 76–79.

Yuceer, M.C.; Hodges, J.D.; Land, S.B., Jr.; Friend, A.L. 1998.

Upland vs. bottomland seed sources of cherrybark oak. In:

Waldrop, Thomas A., ed. Proceedings of the ninth biennial

southern silvicultural research conference; 1997 February 25–

27; Clemson, SC. Gen. Tech. Rep. SRS–20. Asheville, NC: U.S.

Department of Agriculture, Forest Service, Southern Research

Station: 215–221.
87

Horry County, in the Coastal Plain of South Carolina, is

approximately 50 percent wetlands. The Waccamaw Region

(Horry, Georgetown, and Williamsburg Counties) has

experienced a 58-percent population increase during the

period from 1960 to 1990. Population growth trends suggest

that from 1990 to 2020, the total daily population will

increase by 125 percent, representing an additional 510,000

people (Horry County Comprehensive Plan, 1998). The

impact of this on the remaining wetlands will be substantial.

In addition to this problem, S.A.T. scores of South Carolina

students, whereas certainly not the only measure of

academic achievement, do consistently rank below national

averages (South Carolina Dept. of Educ., 2000). For these

reasons, it is necessary to create new enthusiasm for

science education in general and for the value of wetlands in

particular.

In an effort to address this issue, Coastal Carolina University

(CCU) has initiated three programs. In 1998, the School of

Science received an award for Integration of Research and

Education from the National Science Foundation (NSF). This

has provided funding for two of these programs. The first is a

cooperative effort with the Playcard Environmental

Education Center (PEEC), a local nonprofit organization run

mostly by Horry County School District and located along a

blackwater swamp. The grant has provided funding for me to

work with science education students to assist with

educational programs and conduct wetland research. It also

funds minigrants for local high school teachers to bring their

classes to PEEC for research.

The second NSF-funded program is The River Project,

operated by CCU’s Environmental Quality Lab in the Center

for Marine and Wetland Studies. All of our local high schools

are within a 5-minute drive of a river, and, in this program,

they monitor water quality with the assistance of CCU faculty

and students. They report their data on a university Web site

and discuss their findings at an annual student congress.

Funding from Wal-Mart’s Clean Air/Clean Water Program

also supports this program. The Web address is: http://

www.coastal.edu/nsf-aire/river.htm.

Finally, with assistance from the NSF-sponsored State

Systemic Initiative, CCU offered a graduate course last

summer entitled “Wetland Ecology for Teachers”. This was

an activity oriented course designed to give teachers the

knowledge to incorporate wetland education into their

classes in a way that meets the new South Carolina State

Science Education Standards. We also took field trips to

local wetlands, including PEEC, in order to show teachers

how to take advantage of these environments to make

science more alive and interesting for their students. Mostly

middle and high school teachers attended the course. I am

currently researching the extent to which the teachers were

able to incorporate wetlands into their curriculum this year.

The Web address for this course is: http://www.coastal.edu/

science/biology/sgilman/bio778.htm.

ACKNOWLEGMENTS

Thanks to the authors of the NSF-AIRE Grant, Drs. John

Idoux and Val Dunham, the Directors of CCU’s

Environmental Quality Lab, Drs. Susan Libes and Kim

Weaver, Student Coordinator of The River Project, Mr.

Matthew Cline, and the Director of the Waccamaw Math and

Science Hub, Ms. Pam Barnhill.

WETLAND EDUCATION THROUGH COOPERATIVE PROGRAMS BETWEEN

COASTAL CAROLINA UNIVERSITY AND HORRY COUNTY PUBLIC SCHOOLS

Sharon L. Gilman1

1 Instructor of Biology, Department of Biology, Coastal Carolina University, Conway, SC 29528.

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability of

wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.
88

INTRODUCTION

Wetlands and agriculture are closely linked together, dating

back to the prehistoric period. Both wetlands and agriculture

have greatly influenced humankind. Available evidence

suggests that human settlements started in and around the

wetlands (Williams 1990). Long before humans learned to

grow food, they depended, at least partly, on wetlands for

their sustenance. Agriculture had its beginning in the

wetlands and grew at the expense of wetlands (and forests).

Today, about 12 percent of the earth’s land area is under

agriculture. Wetlands occupy < 6 percent of the land area. Of

this, rice culture accounts for about 15 percent and provides

staple food for about 40 percent of the world’s human

population (Hook 1993).

The concern for wetlands is rather recent—only about three

decades old. Until recently, wetlands were treated with

contempt as wastelands, worthy of drainage and reclamation

for agriculture and other land uses. Now that many functions

and values of wetlands have been recognized (Mitsch and

Gosselink 1993), efforts are made to conserve and restore

them. Wetlands, which have been the victim of agriculture,

are also projected today as protectors of water resources

against the impacts of agriculture. In this paper, I try to trace

the linkages between wetlands and agriculture through the

millennia and show how agriculture continues to impact

upon the wetlands. I conclude with a call for an appropriate

policy to strike a balance between agriculture and wetland

conservation.

AGRICULTURE IN WETLANDS

Agriculture is deeply rooted in wetlands. Since prehistoric

times, agriculture has been practiced in and sustained by

wetlands. Agriculture is known to have started in the Middle

East where seasonally flooded wetlands—the riverine

floodplains—provided the ideal environment. The crops

required no subsidy of energy and nutrients. Seeds were

sown after the floods receded exposing the wet soils

enriched with fresh sediments, and the decaying terrestrial

vegetation provided the nutrients. Such low-subsidy

agriculture in floodplains continues today in many parts of

the world. In India, the local communities raise a variety of

crops, such as cucurbits, chillies, tomatoes, etc., not only in

the riparian fringes, but even on the riverbed. The seasonal

wetlands throughout the arid and semi-arid regions of the

world are widely used for raising a variety of crops during

the low water period; for example, dambos are valuable

habitats for agriculture and grazing in Zimbabwe (Scoones

and Cousins 1994). Such agriculture relies heavily on the

timing, frequency, and intensity of flood events.

As the human societies depended increasingly more on

agriculture and developed settlements around wetlands,

agriculture was diversified greatly to exploit many kinds of

wetlands. Humans also recognized the food value of many of

the wetland plants, which were cultivated in the wetlands by

excluding other competitors. Rice (including deep-water rice)

and taro are well-known examples. Trapa bispinosa,

Nelumbo nucifera, Euryale ferox, several sedges (Cyperus

esculentus) and grasses (Panicum, Echinochloa, etc.) are

cultivated extensively in seasonal wetlands of South and

Southeast Asia for food.

At the same time, in several parts of the world, agriculture

expanded further into the waterbodies. Deep-water rice, for

example, is grown in India, Bangladesh, Myanmar, Thailand,

Vietnam, Mali, and Niger in waters where depths may

exceed 6 m (Vergara 1992). In shallow marshes and littorals

of large lakes, rapid vegetative growth of certain

macrophytes, such as Vossia, Cyperus papyrus, Salvinia

spp., Eichhornia crassipes, Echinochloa, Paspalum, etc.,

produces huge biomass, which decays very slowly resulting

in the accumulation of thick layers of detritus mixed with

mineral matter. This decaying mass of organic matter

becomes afloat with the rise in water level following the rains.

These floating islands also become colonized with a variety

of terrestrial and wetland plants and often grow in size with

further accumulation of organic and mineral matter. Floating

islands occur widely throughout the tropics and are known

WETLANDS AND AGRICULTURE:

ARE WE HEADING FOR CONFRONTATION OR CONSERVATION?

Brij Gopal1

1 Associate Professor, School of Environmental Sciences, Jawaharlal Nehru University, New Delhi 110067, India.

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability of

wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.

Abstract—Wetlands and agriculture are closely linked. Historically, agriculture had its beginning in riparian wetland habitats

and expanded into other wetlands. Later, large areas of riverine, palustrine, and coastal wetlands were converted into paddy

fields or drained for agriculture. Agriculture has grown most at the expense of natural wetlands. Today, the intensive

agriculture depends upon heavy inputs of water (irrigation) and agrochemicals. Thus, agriculture now threatens the

remaining wetlands through alteration of hydrological regimes, siltation, and pollutants. Conservation of wetlands requires

an integrated, balanced, and coordinated approach to the management of water resources whereby the impacts of

agriculture on wetlands are minimized without compromising agricultural production.
89

variously as camalotes in Brazil, embalsados in Argentina,

Batumes in Paraguay, sudds in Africa, and phumdis in

eastern India (John 1986, Por 1995, Tombi Singh and Singh

1995). They are known also from outside the tropics; e.g., in

the Mississippi River Delta in Louisiana (Sasser and others

1995) and in Lake Posta Fibreno in Italy (V. Chet, personal

communication). Probably taking cue from these natural

floating islands, and apparently as a response to the

uncertainty of availability of water or to the threats from

flood, and recognizing the freedom of moving them around

at their will, humans started cultivating upon these floating

islands. The chinampas (floating gardens) in Mexico are a

well-known example (Coe 1964, West and Armillas 1950). In

China, rice cultivation on artificial floating islands in lakes

and rivers is known to have been practiced for more than

1,600 years (You Xiuling 1991). Some of the floating islands

on the Yangtze River are reported to have been so large that

several families lived on them together with their pets. The

practice of creating floating islands from lake mud, decaying

organic matter, and macrophytes and cultivating vegetables

on them continues even today in several parts of South and

Southeast Asia. Elaborate descriptions of these floating

islands are available for Malaysia (Department of Agriculture

1939), New Guinea (Serpenti 1965), Myanmar (Annandale

1918), and Kashmir in India (Kaul and Zutshi 1966,

Moorcroft and Trebeck 1841, Sahni 1927, Sturtevant 1970).

CONVERSION OF WETLANDS FOR AGRICULTURE

The most common form of agriculture in wetlands is,

however, paddy cultivation. Evidence of rice culture dates

back to the earliest age of humans (Hook 1993).

Domestication of rice started in shallow swamps (Chang

1976, Harlan 1977), and probably independently in China,

Thailand, and India (Gorman 1977, Ho 1977, Vishnu-Mittre

1977). With the growing demand for food, seasonal marshes

throughout South and Southeast Asia and in China were

modified into paddy fields as man-managed wetlands. It was

only after the European colonization in the 18th and 19th

centuries that the coastal swamps and mangroves were

rapidly converted to paddy fields (Richards 1990). The

conversion of extensive mangroves (Sunderbans) in the

Ganga-Brahmaputra Delta started at the end of the 18th

century by then the East India Company, which required the

private landowners to clear and reclaim Sunderban forests

and swamps for rice cultivation. By the 1870s, 2790 km2 of

mangroves had been converted to rice fields. The pace of

conversion got further accelerated, and though enormous

deposits of sediments transported by the two rivers had

resulted in the expansion of Sunderbans, 2,750 km2 of

Sunderbans were reclaimed between 1880 and 1940 and

another 5,230 km2 in the next 40 years (Richards 1990).

Similarly, the conversion of mangrove swamps in the

Irrawady River Delta started after the British began

colonizing Burma in 1852. During the period from 1880 to

1920, the area under rice grew from 3,979 km2 to 12,059

km2, and wetlands declined from 9,059 km2 to 4,698 km2.

The story of conversion of wetlands including mangroves in

the Chao Phraya Delta under the British colonial rule is also

similar, though it required a great deal of water management

and flow regulation. The independence of nations in the

region in the middle of the 20th century did not help check

the loss of wetlands, particularly the mangroves. The rapid

growth of human population and the need for self-sufficiency

in agricultural production led to expansion of cultivated land

area and intensification of agriculture. This resulted in further

decline of mangroves in the deltas of all major rivers in the

Indian subcontinent. It is estimated that recent losses of

mangroves have been 6 percent in Indonesia, 8 percent in

Malaysia, 20 percent in Thailand, and 50 percent in the

Philippines (Gosselink and Maltby 1990). Mangroves in the

Mekong River Delta in Vietnam suffered extensive

destruction (estimated 24 000 ha) by defoliants during the

war, and, later, they have been rapidly converted to paddy

fields and aquaculture (Khoa and Roth-Nelson 1994).

Other inland marshes and swamps have also been

extensively converted to paddy fields and fishponds. In

India, the fisheries department promoted conversion of

marshes into fishponds throughout the Gangetic Plains

(Jhingran 1992). A recent example is that of Lake Kolleru—a

shallow lake covering 900 km2 between the floodplains of

River Krishna and Godavari. More than 90 percent of the

lake has been converted into large fishponds—an activity

promoted by the government to increase fish production

(Gopal 1991).

Further losses of wetlands have occurred throughout the

tropics for agriculture in the floodplains. The rivers have

been regulated extensively by levees and dykes to isolate

the floodplains. There are no published estimates of the loss

of floodplain wetlands. However, it must be pointed out that

the growing needs of water for irrigated agriculture have also

resulted in an equally large area of man-made wetlands in

the form of tens of thousands of irrigation tanks and

reservoirs (Gopal and Krishnamurthy 1993). Vast areas of

wetlands have also developed by waterlogging due to

seepage (or rise in water table) from the irrigation network.

In the temperate regions, however, the wetlands have been

extensively drained and converted into terrestrial systems. It

is estimated that over 1.6 million km2

 of wetlands had been

drained until 1985 (L’vovich and White 1990) of which threefourths were drained in the temperate regions. Williams

(1990) and Gosselink and Maltby (1990) have discussed in

detail the wetland drainage for agriculture with detailed

examples from U.S., Europe, and Australia. The extent and

history of wetland reclamation in Holland is indeed

impressive. People used coastal salt marshes as early as

500 B.C. Reclamation with polders proceeded with

increasing pace until around the 14th century, and after 1500,

agriculture fuelled the reclamation at a much faster rate. The

limited land space motivated a larger plan in 1918 to reclaim

Zuider Zee to add 2,050 km2 of agricultural land. The

drainage was accompanied by extensive peat mining and

elimination of peat bogs. In the Netherlands alone, 20,000

km2

 (two-thirds of the present-day Netherlands) of wetlands

have been reclaimed from the sea, freshwater lakes, lowlying riverine silts, and peat areas. Similar drainage of

peatlands occurred in other parts of Europe much earlier.

Drainage has been far more extensive in the United States,

though it started only in the mid-19th century. As early as

1921, the U.S. Department of Agriculture had identified

about 37 million ha of wetlands that were “in need of

drainage” in the eastern parts of the country. The wetlands in

the presettlement United States are estimated to cover
90

74.87 million ha (OTA 1984) of which more than 50 percent

had been lost until 1975. More than 80 percent of the total

wetland loss has occurred due to conversion to agriculture.

Extensive drainage has occurred in bottomland forests and

wet prairie regions where a 13.7-percent wetland loss

occurred between the 1950s and 1970s (Gosselink and

Maltby 1990, Williams 1990). The riparian and coastal

wetlands and peatlands have also been lost similarly though

to a much lesser extent. Wetland loss to agriculture by

conversion continues even today. Bernert and others (1999)

report that in the Willamette Valley (Oregon) about 3800 ha

(2.1 percent of the wetlands in the 1980s) of wetlands had

been lost to uplands during 1980 to 1994, of which a 70-

percent loss was associated with agriculture and only 6

percent was lost to urbanization.

IMPACTS OF WETLAND CONVERSION

Conversion of wetlands to agriculture means more than just

a loss. The resulting land use changes as well as the

agricultural practices in and around the remaining wetlands,

which have many impacts upon the wetlands. Several recent

studies have shown that the conversion of wetlands to

agriculture or other land uses (including forestry) impacts

upon the biodiversity though various taxonomic groups

respond differently and at different spatial scales. Mensing

and others (1998) have observed that in the riparian

wetlands of the Northern temperate United States, shrub

carr vegetation, amphibians, and birds are influenced by

land use at relatively smaller scales (500 and 1000 m),

whereas fish respond to land use at landscape level (2500

m or more). Diversity and richness of shrub carr vegetation,

birds, and fish generally decrease with increasing cultivation

in the landscape. A decrease in the proportion of open

water to rangeland results in an increase of amphibian

abundance but a decline of fish abundance. They also

reported that wet meadow vegetation, aquatic macroinvertebrates, amphibians, and fish respond to local

disturbances. Galatowitsch and others (1999) examined

changes in floristic composition corresponding to land use

differences at site to landscape levels in wet meadows

associated with prairie glacial marshes in Minnesota. They

observed that under the impact of agriculture, together with

urbanization, the vegetation composition shifted from native

graminoid and herbaceous perennial abundance to annuals

or introduced perennials. Bethke and Nudds (1995), who

analyzed data on duck abundance in Canadian prairieparklands or the period 1975–89, observed that the recent

decline in the number of breeding ducks, particularly in the

West, have been partly due to loss of habitat to agriculture,

in addition to loss to climatic change (drought). Agriculture

can impact on aquatic invertebrates also in temporary

wetlands. Based on a study of the resting eggs, shells, and

cases remaining after wetlands dried in the prairie pothole

region, Euliss and Mushet (1999) observed that the aquatic

invertebrates were negatively impacted by intensive

agriculture. There were more taxa and greater numbers of

cladoceran resting eggs (ephippia), planorbid, and physid

snail shells, and ostracod shells in wetlands within

grasslands than in croplands.

The changes in agricultural wetlands, such as paddy fields,

depend upon the agricultural practices involving removal of

plants other than the crop, water management, and the use

of agrochemicals. In less intensive paddy cultivation, the

paddy fields support large biodiversity and their productivity,

taking into account that the production of all consumable

biota is substantially high (Heckman 1979). Under intensive

cultivation, the biodiversity is greatly reduced. Aquaculture is

another form of agricultural activity that has adversely

impacted the coastal wetlands in many countries through

loss of biodiversity and pollution.

The agricultural impacts on wetlands are far more complex.

The requirements of water for irrigation directly impinge on

the wetlands as water flows are regulated and diverted.

Shallow and smaller wetlands in drier climates are more

severely affected as their water is used for irrigation. In

Greece, irrigation is reported to be the most important

activity negatively influencing all functions and values of

Ramsar wetlands (Gerakis and Kalburtji 1998). Diversion of

water and reduced freshwater flows to estuarine areas have

affected the mangroves worldwide by way of changes in

species composition due to increased salinity. The fertilizers

and pesticides applied in the field find their way through

surface runoff and subsurface flow into adjacent wetlands.

Furthermore, the agricultural activities enhance erosion

resulting in increased input of sediments into the wetlands.

Thus, changes in the hydrological regimes and an increase

in the sediment and nutrient loading impact upon the biota

and ecosystem processes in wetlands.

Zalidis and others (1997) identified the four most frequent

factors that caused change in the ecological character of

Greek wetlands. Of these, agricultural and municipal

pollution, causing changes in water quality, accounted for 54

percent, construction of irrigation schemes and diversion of

water courses, for 12 percent, and the expansion of

agriculture and settlement for 32 percent of the damage to

wetlands. Change in water regime affected 50 percent of the

springs and 40 percent of the rivers; loss of wetland area

affected 60 percent of the marshes and 52 percent of the

estuaries, whereas all deltas and 75 percent of rivers had

their water quality impacted. Wetlands in the Evros River

Delta are also affected by pollutants transported by the river

from agricultural area in the catchment lying in Bulgaria,

Turkey, and Greece (Angelidis and Athanasiadis 1995).

In Southern Ontario (Canada), cumulative effects of

agricultural land drainage (runoff, subsurface flow, and

nutrient loss) have been estimated to account for the loss of

47 percent of wetlands from 1800 to 1990 (Spaling 1995). In

northern prairie wetlands in North Dakota, Freeland and

others (1999) observed higher sedimentation and

fertilization rates in wetlands next to cultivated fields as

indicated by higher phosphorus, organic matter, and nitrateN concentrations in subsoils (15–60 cm) of wetlands

surrounded by cultivated land than in those surrounded by

grasslands.

In a study of Dives marshes (France), Granval and others

(1993), however, observed that agricultural practices such

as fertilizer use and grazing are beneficial to earthworms,

and, therefore, to their various predators such as snipes

(Gallinago gallinago). The biomass of earthworms in grazed

meadows was more than 10 times higher than in the

adjacent reedbed (Arundo phragmites).
91

Data on agricultural impacts on wetlands in the tropics are

rare. Kassenga (1997) identified irrational use for agriculture

and pollution to be responsible for the degradation and loss

of wetlands in the basin of Lake Victoria (Tanzania). He also

reported subsidence of wetlands due to excessive extraction

of water. In India, practically all wetlands are affected by

agriculture though there are no specific data. Pesticides in

the fields surrounding the Keoladeo National Park

(Bharatpur) are considered to be a major threat because the

park depends entirely on water from outside (Vijayan 1995).

In Kerala (South India), intensive use of fertilizers for rice

cultivation has caused widespread eutrophication of

backwaters.

WETLANDS AS REGULATORS OF NUTRIENT FLUX

Natural wetlands, which lie at the interface between

agricultural uplands and the deep open waters, act as

recipients of sediments and agrochemicals and are known to

regulate their flux to the lakes and rivers. As pointed out

above, the agrochemicals entering the wetlands are

recognized to be among the factors causing degradation of

wetlands. Yet, these very wetlands are valued for the same

functions of intercepting nutrients and other substances, and,

hence, for protecting the downstream waters (Lowrance and

others 1984, Schlosser and Karr 1981, Weller and others

1994, Whigham and others 1988). Hillbricht-Ilkowska and

Kostrzewska-Szlakowska (1993) have shown that the lake

littorals are more efficient in removing nutrients (particularly

nitrogen) than the riparian zones. However, there is growing

realization that the natural wetlands will be quite limited in

absorbing agricultural wastewater (Peterson 1998). Natural

wetlands have in most places either been lost or are highly

degraded. Furthermore, the natural wetlands in the tropics,

e.g., in the Indian subcontinent, also appear to receive

wastewaters and runoff from nonpoint sources (including

agriculture) far beyond their buffering capacity. In this context,

the constructed wetlands technology is being extended to

intercept agricultural runoff and remove pollutants from it

(Hammer 1992, Kern and Idler 1999, Peterson 1998,

Rodgers and Dunn 1992). Thus, constructed wetlands may

replace the natural wetlands and become partners with

agriculture in integrated water and nutrient management.

CONCLUSION

The continuing loss of natural wetlands to agriculture and

the increasing emphasis on intensive agriculture, which

causes wetland degradation, raise doubts over the future of

wetlands, including agricultural wetlands. Can the man-made

wetlands replace the functions and values of natural

wetlands? Can wetlands be really conserved without

checking agricultural impacts? The sustainability of wetlands

and agriculture is interlinked with that of the water resources.

The sustainability of agro-ecosystems also depends upon

wetlands because the latter provide, besides irrigation water,

also crop pollinators, some frost protection, and predators of

crop pests (Gerakis and Kalburtji 1998). The future scenario

appears to be worse. Discussing the loss of wetlands to

drainage in Eastern Europe, Hartig and others (1997) have

pointed out that following climate change, higher

temperatures and greater evapotranspiration may alter the

hydrologic regime such that freshwater wetlands are further

encroached upon by agricultural land use. Wetland

conservation seemingly faces confrontation with agriculture.

The problem lies in the national policies and their

implementation. In recent years, several countries have

adopted national wetland policies, but, only rarely,

agriculture is recognized as a major sector competing with

wetlands for water or agriculture leading to degradation of

wetlands. There has been no effort to resolve the conflict

with the agricultural policy. Within the United States, Nelson

(1986a, 1990) had discussed various aspects of the

agricultural policy in some detail. He points out the policy

crisis results from contradictory incentives of central

government to both drain and preserve wetlands. A similar

situation exists also in the United Kingdom (Nelson 1986b).

In Greece, Zalidis and others (1997) pointed out the

inefficacy of the agricultural policy, which could not protect

small wetlands or deal with the loss of wetlands from

intensified use of existing farmland. However, Pyrovetsi and

Daoutopoulos (1998) observed that the wetland farmers had

a more negative attitude toward the wetland resources and

were more ignorant of conservation issues or the impact of

their practices on the environment than plain farmers.

Most of the developing countries, with great population

pressures, are compelled to focus on increasing agricultural

production despite declining returns from the degraded land

resources. In the process, there are ever-increasing

demands on water resources, which have depleted and

degraded rapidly. Both agriculture and wetlands are

threatened. Therefore, the need for an integrated, balanced,

and coordinated approach to water resource management is

greater than ever before. We need an integrated policy,

which takes care of both natural wetlands and agriculture so

as to minimize the impacts of agriculture on wetlands

without compromising agricultural production.

ACKNOWLEDGMENTS

I am grateful to Dr. Marjorie Holland and Dr. Charles M.

Cooper for their invitation to the conference, and to the Croft

Institute for International Studies at the University of

Mississippi and the Society of Wetland Scientists for

providing necessary financial support for my participation. I

thank Professor D.P. Zutshi and Mr. V. Chet for valuable

information on the cultivation of floating islands.

REFERENCES

Angelidis, M.O.; Athanasiadis, A.I. 1995. Pollution mechanisms. In:

A Ramsar wetland—delta of the River Evros, Greece. Aquatic

Sciences. 57(2): 161–171.

Annandale, N. 1918. Aquatic molluscs of the Inle Lake and

connected water. Records of the Indian Museum. 14: 103–182.

Bernert, Joseph A.; Eilers, Joseph M.; Eilers, Benn J. [and others].

1999. Recent wetlands trends (1981/82–1994) in the Willamette

Valley, Oregon, U.S.A. Wetlands. 19(3): 545–559.

Bethke, R.W.; Nudds, T.D. 1995. Effects of climate change and land

use on duck abundance in Canadian Prairie–Parklands.

Ecological Applications. 5(3): 588–600.

Chang, T.T. 1976. Rice: Oryza sativa and Oryza glaberrima

(Graminae: Oryzae). In: Simmonds, N.W., ed. Evolution of crop

plants. London: Longman: 98–104.
92

Coe, Michael D. 1964. The chinampas of Mexico. Scientific

American. 211 (1): 90–98.

Department of Agriculture (Economics Branch). 1939. Padi planting

methods in Malaya. Malayan Agricultural Journal. 27(2): 40–59.

Euliss, Ned H., Jr.; Mushet, David M. 1999. Influence of agriculture

on aquatic invertebrate communities of temporary wetlands in

the prairie pothole region of North Dakota, U.S.A. Wetlands.

19(3): 578–583.

Freeland, John A.; Richardson, Jim L.; Foss, Lynn A. 1999. Soil

indicators of agricultural impacts on northern prairie wetlands:

Cottonwood Lake research area, North Dakota, U.S.A. Wetlands.

19(1): 56–64.

Galatowitsch, S.M.; Whited, D.C.; Lehtinen, R. [and others]. The

vegetation of wet meadows in relation to their land use.

Environmental Monitoring & Assessment. 60(2): 121–144.

Gerakis, Argyrios; Kalburtji, Kiriaki. 1998. Agricultural activities

affecting the functions and values of Ramsar wetland sites of

Greece. Agriculture Ecosystems & Environment. 70(2–3): 119–

128.

Gopal, B. 1991. Wetland (mis)management by keeping people out:

two examples from India. Landscape and Urban Planning. 20:

53–59.

Gopal, B.; Krishnamurthy, K. 1993. Wetlands of South Asia. In:

Whigham, D.F.; Dykyjova, D.; Hejny, S., eds. Wetlands of the

world. I. Inventory, ecology and management. Handbook of

vegetation science. Dordrecht, Kluwer: 15/2: 345–414.

Gorman, C. 1977. A priori models and Thai prehistory: a reconsideration of the beginnings of agriculture in Southeastern Asia. In:

Reed, C.A., ed. Origins of agriculture. The Hague: Mouton: 321–

355.

Gosselink, J.G.; Maltby, E. 1990. Wetland losses and gains. In:

Williams, M., ed. Wetlands: a threatened landscape. Oxford,

U.K.: Blackwells: 296–322.

Granval, P.; Aliaga, R.; Soto, P. 1993. The impact of agricultural

management on earthworms (Lumbricidae), common snipe

(Gallinago gallinago), and the environmental value of grasslands

in the Dives marshes (Calvados); Gibier Faune Sauvage.

10(March): 59–73. In French.

Hammer, D.A. 1992. Designing constructed wetland systems to treat

agricultural nonpoint source pollution. Ecological Engineering. 1:

49–82.

Harlan, J.R. 1977. The origins of cereal agriculture in the Old World.

In: Reed, C.A., ed. Origins of agriculture. The Hague: Mouton:

357–383.

Hartig, Ellen K.; Grozev, Ognyan; Rosenzweig, Cynthia. 1997.

Climate change, agriculture, and wetlands in Eastern Europe:

vulnerability, adaptation, and policy. Climatic Change. 36(1–2):

107–121.

Heckman, C.W. 1979. Rice field ecology in northeastern Thailand.

The Hague: W. Junk, bv. 228 p.

Hillbricht-Ilkowska, A.; Kostrzewska-Szlakowska, I. 1993. Transport

of dissolved forms of nitrogen derived from arable land along the

river-lake system in a post-glacial landscape. In: Gopal, B.;

Hillbricht-Ilkowska, A.; Wetzel, R.G., eds. Wetlands and ecotones:

studies on land-water interactions. New Delhi: National Institute

of Ecology and International Scientific Publications: 61–74.

Ho, P.T. 1977. The indigenous origins of Chinese agriculture. In:

Reed, C.A., ed. Origins of agriculture. The Hague: Mouton: 413–

484.

Hook, Donald. 1993. Wetlands: history, current status, and future.

Environmental Toxicology & Chemistry. 12(12): 2157–2166.

Jhingran, V.G. 1992. Fish and fisheries of India. 3rd ed. New Delhi:

Hindustan Publishing Corp.

John, D.M. 1986. The inland waters of tropical West Africa. Archiv

fuer Hydrobiologie, Ergebnisse der Limnologie. 23: 1–244.

Kassenga G.R. 1997. A descriptive assessment of the wetlands of

the Lake Victoria Basin in Tanzania. Resources Conservation &

Recycling. 20(2): 127–141.

Kaul, V.; Zutshi, D.P. 1966. Some ecological consideration of floating

islands in Srinagar lakes. Proceedings of the National Academy

of Sciences of India; 36B: 273–281.

Kern, J.; Idler, C. 1999. Treatment of domestic and agricultural

wastewater by reed bed systems. Ecological Engineering. 12:

13–25.

Khoa, L.V.; Roth-Nelson, W. 1994. Sustainable wetland use for

agriculture in the Mekong River Delta of Vietnam. In: Mitsch,

W.J., ed. Global wetlands: Old World and new. Amsterdam:

Elsevier: 737–748.

L’Vovich, M.I.; White, G.F. 1990. Water. In: Burner, B.L., III; Clark,

W.C.; Kates, R.W. [and others], eds. The earth as transformed by

human action. New York: Cambridge University Press.

Lowrance, R.; Todd, R.; Fail, J. [and others]. 1984. Riparian forests

as nutrient filters in agricultural watersheds. Bioscience. 34: 374–

377.

Mensing, D.M.; Galatowitsch, S.M.; Tester, J.R. 1998. Anthropogenic

effects on the biodiversity of riparian wetlands of a northern

temperate landscape. Journal of Environmental Management.

53(4): 349–377.

Mitsch, W.J.; Gosselink, J.G. 1993. Wetlands. 2nd ed. New York: Van

Nostrand-Reinhold. 722 p.

Moorcroft, William; Trebeck, G. 1841. Travels in the Himalayan

provinces of Hindustan and the Panjab; in Ladakh and Kashmir.

London: John Murray.

Nelson, R.W. 1986a. Wetland management strategies balancing

agriculture with conservation: a national survey of expert opinion.

Environmental Professional. 8(1): 54–69.

Nelson, R.W. 1986b. Wetlands policy crisis U.S.A and U.K.

Agriculture Ecosystems & Environment. 18(2): 95–122.

Nelson, R.W. 1990. Can experts agree on policy directions toward

wetlands and agriculture. Environmental Professional. 12(2):

131–143.

OTA (Office of Technology Assessment). 1984. Wetlands: their use

and regulation. Washington, DC: U.S. Government Printing

Office.

Peterson, H.G. 1998. Use of constructed wetlands to process

agricultural wastewater. Canadian Journal of Plant Science.

78(2): 199–210.

Por, F.D. 1995. The pantanal of Mato Grosso (Brazil): world’s largest

wetlands. Monographiae Biologicae 73. Dordrecht: Kluwer. 122 p.
93

Pyrovetsi, M.; Daoutopoulos, G. 1999. Farmers’ needs for nature

conservation education in Greece. Journal of Environmental

Management. 56(2): 147–157.

Richards, J.F. 1990. Agricultural impacts in tropical wetlands: rice

paddies for mangroves in South and Southeast Asia. In:

Williams, M., ed. Wetlands: a threatened landscape. Oxford,

U.K.: Blackwells: 217–233.

Rodgers, J.H., Jr.; Dunn, A. 1992. Developing design guidelines for

constructed wetlands to remove pesticides from agricultural

runoff. Ecological Engineering. 1: 83–95.

Sahni, B. 1927. A note on the floating islands and vegetation of

Khajiar, near Chamba in the N.W. Himalayas. Journal of the

Indian Botanical Society. 6: 1–7.

Sasser, C.E.; Gosselink, J.G.; Swanson, E.M.; Evers, D.E. 1995.

Hydrologic, vegetation and substrate characteristics of floating

marshes in sediment rich wetlands of the Mississippi River Delta

Plain, Louisiana, U.S.A. Wetlands Ecology and Management. 3:

171–187.

Schlosser, I.J.; Karr, J.R. 1981. Riparian vegetation and channel

morphology impact on spatial patterns of water quality in

agricultural watersheds. Environmental Management. 5: 233–243.

Scoones, I.; Cousins, B. 1994. Struggle for control over wetland

resources in Zimbabwe. Society & Natural Resources. 7(6): 579–

594.

Serpenti, L.M. 1965. Cultivators in the swamps: social structure and

horticulture in a New Guinea society (Frederik-Hendrik Island

West New Guinea). Assen, Netherlands: Van Gorcum-Prakke.

Spaling, H. 1995. Analyzing cumulative environmental effects of

agricultural land drainage in Southern Ontario, Canada.

Agriculture Ecosystems & Environment. 53(3): 279–292.

Sturtevant, W.C. 1970. Agriculture on artificial islands in Burma and

elsewhere. Eighth International Congress of Anthropological and

Ethnological Sciences; Tokyo and Kyoto; 1968: 11–13. Vol. 3.

Tombi Singh, H.; Singh, R.K.S. 1995. Loktak Lake, Manipur. WWFIndia, New Delhi. 69 p.

Vergara, B.S. 1992. Tropical wet rice systems. In: Pearson, C.J., ed.

Field crop ecosystem. Amsterdam: Elsevier: 167–182.

Vijayan, L. 1995. Keoladeo National Park. WWF-India, New Delhi. 77

p.

Vishnu-Mittre. 1977. Changing economy in ancient India. In: Reed,

C.A., ed. Origins of agriculture. The Hague: Mouton: 570–588.

Weller, D.E.; Correll, D.L.; Jordan, T.E. 1994. Denitrification in

riparian forests receiving agricultural discharges. In: Mitsch, W.J.,

ed. Global wetlands: Old World and new. Amsterdam: Elsevier:

117–131.

West, Robert C.; Armillas, P. 1950. Las chinampas de México:

Poesia y realidad de los “jardines flotantes.” Cuadernos

Americanos. Año 9, No. 2: 165–182. Vol. 50.

Whigham, D.F.; Chitterling, C.; Palmer, B. 1988. Impacts of freshwater wetlands on water quality: a landscape perspective. Environmental Management. 12: 663–671.

Williams, Michael, ed. 1990. Wetlands: a threatened landscape.

Oxford, U.K.: Blackwells. 419 p.

You Xiuling. 1991. Rice planting on the water surface in ancient

China. Chinese Journal of Rice Science. 5.2: 79–82. In Chinese.

Zalidis, George C.; Mantzavelas, Antonis L.; Gourvelou, Eleni. 1997.

Environmental impacts on Greek wetlands. Wetlands. 17(3):

339–345.
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INTRODUCTION

Current interest in wetland restoration, enhancement, and

mitigation has led to an increased need for sources of plants

that are well adapted to local environmental and soil/water

conditions. Most plant materials available commercially are

from widely varying geographic regions. Plant materials

originating from areas with vastly different physical and

environmental characteristics are often not well suited to

conditions on the planned planting site, and use of these

plants can lead to less than satisfactory performance.

McNaughton (1966) found significant differences in growth

habit, environmental tolerances, and dormancy patterns of

several ecotypes of three Typha L. species collected from

diverse sites in the United States. Growth patterns of each

ecotype were matched for optimum survival and productivity

to the environmental conditions common to their original

source location. Although the ideal recommendation would

be to use planting materials collected from local ecotypes,

this is often impractical because availability may be lacking

and costs prohibitive (U.S. Army Engineer Waterways

Experiment Station 1978). Therefore, an alternate method of

obtaining adapted plant materials is required.

Awareness of genetically controlled ecotypic differences

between plant populations has led to development of a new

type of plant release within the USDA-Natural Resources

Conservation Service (NRCS) Plant Materials Program

called a source-identified release. A source-identified

release was collected from a natural plant population and

has not undergone any testing or selection prior to its

release.

Staff at the Jamie L. Whitten Plant Materials Center (PMC),

Coffeeville, MS, collected plants of several emergent wetland

species within the State and released three of these for use

in various types of wetland plantings. Although these

ecotypes are best suited to areas within Mississippi,

ecological differences within the State and among adjoining

States are much smaller than in some other parts of the

country. This fact along with the fairly broad natural

geographic range of these species suggests that they may

also be planted successfully in other areas in the

Southeastern United States with comparable climatic,

hydrologic, topographic, and soil features.

Plant material of these releases was provided to a limited

number of local plant producers for increase and sale to the

public. At present, vegetative propagules (clonal material)

are being marketed; however, production of seedlings may

be more desirable because genetic recombination could

lead to greater diversity in the planting population and

increased ability to withstand environmental stresses. This

paper presents descriptive information on each of these

releases as well as results of a study conducted to

determine seed propagation methods for these plants.

PLANT COLLECTIONS

Indian Bayou

Thalia dealbata is a member of the Maranthaceae or

arrowroot family. It is a rare inhabitant of shallow wetlands in

the Mississippi Delta and other Southeastern States. The

Indian Bayou source of powdery thalia was collected by

Travis Salley in 1989 from a homeowner’s yard in Bolivar

County, Mississippi. According to the homeowner, the plants

were originally moved from Washington County, Mississippi.

It is a herbaceous perennial arising from thick (up to 4 cm in

diameter) rhizomes, with a bluish, glaucous coating on

leaves, flower stalks, and flowers. The 1- to 1.2-m long

leaves arise from the base of the plant and have a shape

reminiscent of a Canna L., with a stout petiole and a large,

ovate to lanceolate leaf blade. Flowers are produced from

late May to September with fruit maturing throughout the

summer. The attractive purple to bluish flower clusters are

produced on scapes extending 0.6 to 1 m above the foliage.

Fruit type is an urticle, consisting of a bladdery, membranous

covering loosely surrounding a single seed. Seeds are

approximately 6 mm in diameter, subglobose, dark-brown

THREE MISSISSIPPI ECOTYPES OF WETLAND PLANTS

Janet M. Grabowski1

1 Biologist, USDA-NRCS, Jamie L. Whitten Plant Materials Center, 2533 County Road 65, Coffeeville, MS 38922-2652.

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability of

wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.

Abstract—In 1996, the USDA-Natural Resources Conservation Service (NRCS) Jamie L. Whitten Plant Materials Center

(PMC) released three locally collected, source-identified wetland plants. Indian Bayou source powdery thalia (Thalia

dealbata Fraser ex Roscoe) and Leflore source creeping burhead [Echinodorus cordifolius (L.) Griseb.] were collected in

the Mississippi Delta and Leaf River source woolgrass [Scirpus cyperinus (L.) Kunth] from southeastern Mississippi. Both

vegetative propagules and seedlings can be planted. PMC seed-germination studies have shown that Indian Bayou and

Leaf River germinate readily after exposure to cold, moist conditions (stratification). Indian Bayou germination and seedling

growth was best in a moist, not saturated, growing medium, whereas best germination of Leaf River was in a saturated

medium, and seedling growth was better in a moist medium. Germination of Leflore was very poor for all seed treatments in

the test, even though seedlings have established in PMC plant-production ponds.
95

speckled with tan or gray, with a conspicuous tan to brown

hilum. There are approximately 3,184 seeds per kilogram.

Leaf River

Woolgrass is a bulrush, in the Cyperaceae or sedge family

that is commonly found in shallow wetlands throughout the

State of Mississippi. Its native range covers most of the

Eastern United States and Canada, extending westward into

eastern Texas [USDA-Soil Conservation Service (no date)].

Leaf River source was collected by B.B. Billinglsey, Jr.,

Harvey Huffstatler, and Jeff Tillman from a site near the Leaf

River in Jones County, Mississippi. It is a clump-forming

perennial with short rhizomes. The grass-like basal leaves

are up to 1.5 m in length and arch outwards from the base of

the plant like a fountain. The flowering culms are 1.2 to 1.8 m

tall, leafy, somewhat coarse, and obtusely triangular. The

dense inflorescence contains numerous pale-green spikelets

that become brown and wooly as the seeds mature. The tiny

achenes are light tan, with six long, twisting perianth bristles.

Flowering begins in June, and seeds mature by September.

The achenes were too small to be accurately counted by

PMC equipment, but it was estimated that there are upwards

of 18 million per kilogram.

Leflore

Creeping burhead can be found in swamps, marshes, and

ditches in the Southeast and Lower Midwest, westward into

Oklahoma and Texas (Godfrey and Wooten 1979). It is a

member of the Alismataceae or water-plantain family. The

Leflore source was collected by B.B. Billingsley, Jr., Joe

Snider, Joel Douglas, and Janet Grabowski from a flooded

lowland in Leflore County, Mississippi. Leflore is an annual or

short-lived perennial that often spreads or creeps by rooting

scapes. Basal leaves are broadly ovate, cordate at the base,

5 to 18 cm long and almost as wide. The principal veins are

conspicuous and impressed on the upper surface of the leaf

blade. Petioles are 10 to 50 cm long, enlarged and spongy

towards the base. The leafy scapes are upright when young,

becoming prostrate, up to 1 m or more in length, producing

new plantlets at the tip and nodes. Numerous whorls of 12-

to 20-mm wide flowers, with three white petals and rounded,

greenish centers are produced at nodes along the scape.

Flowering begins in June and continues until frost. The

fruiting heads are round, bur-like clusters of small brown,

flattened achenes, which have a long beak on one end.

There are approximately 5 million achenes per kilogram.

VEGETATIVE PROPAGATION

All of these releases can be vegetatively propagated by

division of the parent plant. The large size of the shoots and

rhizomes of Indian Bayou dictate that a fairly large planting

piece is required. A section about 15 to 24 cm long with a

few shoots is ideal. Plantings made by the PMC on a

Wetland Reserve Program (WRP) site have indicated that

Indian Bayou propagules are capable of withstanding fairly

adverse environmental conditions after planting (data not

presented); these results can probably be attributed to the

large amount of food reserves stored in the rhizomes.

Leaf River clumps can be dug and individual shoots divided

somewhat easily. Older clumps tend to die back in the

center, so actively growing shoots around the margins of the

plant should be selected. It is also a good idea to discard

shoots containing an old flowering culm because these

either will not regenerate or, if they do, the shoot produced

will be very weak. Care should be taken in handling Leaf

River plants because the edges of the leaves are very sharp

and can cut deeply into flesh.

Established Leflore plants can be dug and divided into a

small number of propagules. These plants do not have a

very deep root system, which makes them fairly easy to dig.

Small plants produced on the flowering scape can also be

removed and planted.

SEED GERMINATION STUDY

Methods and Materials

Seeds were harvested in late summer to fall of 1995 and

1996 (harvest dates varied between species and between

year of collection). Seeds were cleaned immediately after

collection without drying. Leaf River seeds were loosened

from fruit clusters using a brush machine (Westrup a/s

Slagelse, Denmark) and then hand screened to remove inert

matter. Indian Bayou seeds were rubbed over a roughened

surface to remove the papery fruit coverings. Leflore seeds

required only hand screening to remove small amounts of

trash.

Three storage treatments were tested: (1) dry storage in a

cooler at 13 °C and 45 percent relative humidity; (2) moist

storage in a cooler at 5.5 °C; and, (3) water storage in a

cooler at 5.5 °C. Dry storage was tested on all species. Moist

storage was used for all seedlots, except the 1996 Leaf

River lot. Water storage was tested on the 1995 seedlot of

Leaf River and 1996 Leflore only. Small quantities of seeds

of each species were divided from the main lot and placed in

moist and water storage immediately after cleaning. All

seeds in water storage were put in a nylon-mesh bag and

then submerged in tap water in a glass jar. Water was not

changed during the storage period for the first year of the

test; however, it was changed monthly during the second

year in an attempt to limit algae growth. Moist, stored seeds

of Leaf River and Leflore were placed on a moistened brown

paper towel and placed in a self-sealing plastic bag with

sufficient additional water to maintain high moisture levels

during the storage period. Indian Bayou seeds are too large

to allow good seed contact with paper toweling, so seeds

were stratified in moist sphagnum moss from which they

were easily separated at planting. All remaining seeds were

dried thoroughly and placed in a self-sealing plastic bag for

the dry-storage treatments.

Preplanting treatments applied to dry-stored seeds were a 3-

to 4-month stratification period (Strat.), and for Indian Bayou

seed only, scarification (Scar.) using mechanical means, and

combinations of stratification and scarification. In 1995,

seeds were scarified by placing them in a coffee can lined

with sand paper and tumbling with some gravel added to

increase abrasion. A concern that seeds were not uniformly

scarified using this method required use of an alternate

method in the second year where seeds were individually

rubbed against sandpaper. All stratification treatments were

placed in the cooler during November 1995 and 1996.

Treatments were planted in the greenhouse within a 1-week

period in March 1996 and 1997. Stratified, moist-stored, and
96

water-stored Leaf River and Leflore seeds were allowed to

surface dry during the planting process to facilitate sample

counting. Sample sizes used were 100 seeds of Leaf River

and Leflore for each treatment and 20 seeds of the 1995 lot

and 25 seeds of the 1996 Indian Bayou lot. Germination

containers were 17.8 cm by 13.3 cm by 5.9 cm black plastic

bedding-plant liners. Growing medium used was a 3:1 peat

moss/sand growing medium amended with commercially

recommended quantities of pelletized slow-release fertilizer,

dolomitic lime, Micromax micronutrient fertilizer, and

Aquagro wetting agent. Seeds were sown on the surface of

the growing medium, except Indian Bayou seeds, which

were planted approximately 6 mm deep. Seeds were spread

as uniformly as possible on the medium.

Two germination conditions were tested: (1) moist conditions,

where containers were placed on a normal greenhouse

bench and watered regularly (normal bench); and, (2)

continuously saturated conditions on an ebb and flow

greenhouse bench (flood bench). Water on the ebb and flow

bench was maintained at a depth of 6 to 12 mm except for

short periods of time when the bench was drained and

rinsed to remove alga growth. Experimental design was a

factorial experiment in a randomized complete block with

three replications arranged as a split plot, with germination

condition as the main plot and seed treatment (storage

condition and preplanting seed treatment) as the split plot.

An initial seedling count was made when it was deemed that

a sufficient number of seedlings were present to justify

counting. Two additional counts were made at 3-week

intervals following the initial count. A few seedlings of some

species died before the initial count, especially on the

normal greenhouse bench, where the surface of the growing

medium dried for short periods of time. Dead seedlings were

counted, and their numbers were included in the initial count

because germination had occurred. Later counts included

only those plants that had some green tissue. Total

germination could not be determined, so results presented

are from the count with highest overall number of seedlings.

The study was analyzed with years of testing treated

separately. Data was subjected to an analysis of variance,

and appropriate mean separation was performed using a

least significant difference test (LSD) at P = 0.05. All data on

seedlots are referenced by year of seed collection (1995 and

1996), not year of testing.

RESULTS AND DISCUSSION

Indian Bayou Germination

Indian Bayou appears to have complex germination and

seedling growth requirements. Only a few seedlings have

established naturally in PMC ponds, mainly along the margin

of the pond. Previous attempts at seedling production in the

greenhouse indicated that germination was erratic and slow.

Initial germination percentages in this study were fairly low,

and seeds continued to germinate in small numbers

throughout the test. For the 1995 seedlot, there was a

significant response to seed treatment only (table 1). Overall

germination was better on the normal bench than on the

flood bench. In this environment, seeds responded positively

to moist storage and both stratification treatments. The high

germination percentage for scarified plus stratified seeds

compared to stratified seeds indicate that this seedlot

responded to scarification.

For the 1996 seedlot, there was a significant interaction

between seed treatment and germination condition. This

interaction was probably due to poor germination of all

treatments on the flood bench. Also, germination of drystored seeds was poor on both benches. Germination of

scarified seeds on the normal bench was significantly higher

for this seedlot than any other treatment, which is curious

because of the low percentage for this treatment in the

previous year. The scarification method used for this seedlot

was more vigorous and may have increased response to this

treatment. However, this cannot explain the response of

scarified plus stratified seeds, where germination was not

different than stratified seeds.

In both years, seedlings on the flood bench were more

susceptible to infestation by aphids than those on the normal

bench, indicating that the plants were probably stressed.

This improved germination and growth on the normal bench

appears to corroborate observations of seedling

establishment in drier conditions along the margins of PMC

growing ponds. The response to stratification and moist

storage indicates that Indian Bayou has an internal

dormancy that is overcome by cold, moist treatment.

Germination percentages were slightly higher for dry-stored

seeds that were subsequently stratified than for moist-stored

seeds regardless of treatment year, which suggests

capability for long-term seed storage. Mechanical

scarification, especially when accompanied by stratification,

also appears to be beneficial, perhaps because seeds have

a waxy coating that may prevent water infiltration. A previous

attempt at sulfuric acid scarification on a small quantity of

seeds showed that even a short duration (< 1 minute)

treatment was toxic (Janet Grabowski, unpublished data).

Table 1—Indian Bayou mean germination percentages

for seed exposed to two storage conditions, three

preplanting treatments, and two germination conditions

Storage and Germination

pretreatment condition 1995 1996

 - - - Percent - - -

Dry/strat. Normal bench 35 17

Dry 8 1

Dry/scar. + strat. 25 17

Dry/scar. 5 40

Moist 27 15

Dry/strat. Flood bench 0 3

Dry 0 3

Dry/scar. + strat. 15 8

Dry/scar. 0 9

Moist 3 1

LSD (P = 0.05) 17 13

Strat. = stratification; scar. = scarification; LSD = least significant

difference.
97

However, because mechanical scarification did not

consistently improve germination, it appears that

stratification alone is probably sufficient to promote

germination.

Leaf River Germination

Results for the 1995 and 1996 seedlots were similar for

those treatments common to both years of testing (table 2).

The 1995 lot showed a significant response to seed

treatment. Dry-stored Leaf River seeds appeared to

germinate more quickly after stratification, but germination

percentages were not significantly different from the drystorage treatment alone, which conflicts with Isley (1944)

who found that woolgrass seeds require stratification for 6

months or longer to overcome dormancy. This southern

Mississippi ecotype is adapted to areas with only short

periods of chilling winter temperatures and apparently does

not have a dormancy mechanism that requires stratification.

Garbisch and McIninch (1992) also reported a lack of seed

dormancy. There was a trend towards improved germination

of the 1995 seeds stored in water, but the difference was

significantly higher than only the dry-stored seeds on the

normal bench and the moist-stored seeds on the flood

bench.

Leaf River seeds showed a significant response to

germination conditions in both years of testing. Highest

germination was on the flood bench with saturated

conditions. Seedling mortality on the normal bench was

higher for this species, probably because of the small size of

the seedlings, which made them highly susceptible to

desiccation as the surface of the growing medium dried. This

problem was more pronounced in 1996, probably due to

cooler temperatures and longer periods of cloud cover in

1997. However, later seedling growth in both years was

more vigorous for surviving plants on the normal bench.

Seedlings growing on the flood bench showed evidence of

poor root growth due to a lack of aeration, and there was

extensive algae and slime mold growth on the growing

medium surface, which may have been toxic to the

seedlings. Most of the plants on the flood bench were barely

alive by the third evaluation date. Observations made of

plants growing in the wild indicate that this species requires

fairly wet conditions for germination, but plants become

increasingly tolerant of drying substrates as they grow, with

mature plants possessing a higher level of drought tolerance

than would be anticipated for a wetland plant.

Leflore Germination

Mean germination percentages for all treatments were

extremely low (< 5 percent) in both years (data not

presented). For 1995 and 1996 seedlots, there was a

significant response to germination condition, with best

germination on the flood bench. However, these differences

were too minuscule to support a recommendation on proper

germination and seedling growth environments. Storage

treatments yielded the same results in both years. Although

water storage was added in the second year to see if

germination could be improved over that of moist storage, no

improvement was observed. Water- and moist-stored seeds

showed a slight trend towards improved germination

compared to dry storage. An attempt was made to

mechanically scarify some seeds from the 1996 lot, but

visual inspection after treatment did not indicate that this

treatment had any appreciable effect on seed coverings.

Seed germination in the PMC production ponds has been

noted, mainly along the margins and in shallow water;

however, this test apparently did not provide suitable

conditions for germination.

SUMMARY

These three source-identified releases can potentially be

used in wetland plantings in Mississippi and perhaps other

areas in the Southeast. Because of their specific

germination requirements, direct seeding is probably not

advisable. However, greenhouse seedling production is

practical for Indian Bayou and Leaf River, with stratification

recommended for Indian Bayou seeds. Additional research

would be necessary to determine methods to successfully

propagate Leflore from seed.

REFERENCES

Garbisch, E.W.; McIninch, S. 1992. Seed information for wetland

plant species of the Northeast United States. Restoration and

Management Notes. 10: 85–86.

Godfrey, R.K.; Wooten, J.W. 1979. Aquatic and wetland plants of

Southeastern United States monocotyledons. Athens, GA: The

University of Georgia Press. 712 p.

Isley, D. 1944. A study of conditions that affect the germination of

Scirpus seeds. Memoir 257. Ithaca, NY: Cornell University

Agricultural Experiment Station. 28 p.

McNaughton, S.J. 1966. Ecotype function in the Typha communitytype. Ecological Monographs. 36: 297–325.

U.S. Army Corp of Engineers. 1978. Wetland habitat development

with dredged material: engineering and plant propagation. Tech.

Rep. DS–78–16. Vicksburg, MS: U.S. Army Corp of Engineers,

Waterways Experiment Station. 158 p.

U.S. Department of Agriculture. [n.d.] Midwestern wetland flora field

office guide to plant species. Lincoln, NE: U.S. Department of

Agriculture, Soil Conservation Service. 615 p.

Table 2—Leaf River mean germination percentages for

seed exposed to various seed storage conditions, one

preplanting treatment, and two germination conditions

Storage and Germination

pretreatment condition 1995 1996

 - - - Percent - - -

Dry/strat. Normal bench 19 17

Dry 12 9

Moist 15 —

Water 24 —

Dry/strat. Flood bench 35 61

Dry 41 57

Moist 32 —

Water 44 —

LSD (P = 0.05) 8 23

Strat. = stratification; LSD = least significant difference.
98

INTRODUCTION

Recent assessments of afforestation of agricultural lands in

the Mississippi Alluvial Valley (MAV), particularly the Delta

area of Mississippi, have stressed the importance of quickly

attaining the physical structure and stature of forests

(Schweitzer and others 1997). Implementation of land

management plans designed for neotropical migratory birds

in the Lower MAV (LMAV) (Mueller and others 1995) will

benefit from dependable, rapid afforestation as well. Benefits

of afforestation include provision of habitat for middle and

late successional birds, production of pulpwood, production

of sawtimber wood products, and erosion control, as well as

carbon sequestration and accumulation of soil organic

matter. All these benefits are positively associated with the

speed with which afforestation occurs. Rapid afforestation

implies swift accumulation on the landscape of the physical

structure and stature of forest. It is a means of carbon

sequestration (Cannell 1999b, Chang 1999) and of land

rehabilitation in Amazonia (Bauch and others 1999) and

elsewhere (Harrington 1999). Fast development of vertical

forest structure is implicit in the environmental (Joslin and

Schoenholtz 1997) and economic (Pande and others 1999,

Scholtens 1998) analyses of afforestation.

Afforestation, including rapid afforestation, is assumed to be

beneficial to wildlife (Boyle 1999, Cannell 1999a, Helmer

1999, Weaver and Pelton 1994, Weaver and others 1990,

Willoughby and McDonald 1999). On the other hand, certain

native wildlife and grazing animals can hinder afforestation

efforts (Anderson and Katz 1993, Houston 1991, Niyaz and

others 1999).

Vegetation structure is an important determinant of bird

species occurrence and community composition (DeGraaf

1987, DeGraaf and others 1992, James 1971). Hamel

(1992) associates birds with combinations of vegetation

structure, such as trees, shrubs, and herbaceous vegetation.

Afforestation yields unusual elements in secondary

succession, such as tall cottonwood (Populus L. spp.) trees

and herbaceous vegetation with little woody understory. A

first hypothesis is that the bird community developing in

afforestation will reflect vegetation structure.

Wintering birds use the early successional herbaceous

community. During rapid afforestation, the early successional

periods are shorter than during natural succession. Our

second hypothesis is that early successional species may

not benefit from rapid afforestation of agricultural fields as

much as from natural succession.

To examine the assumptions that rapid afforestation is

beneficial to wildlife and that bird species occurrence is a

function of vegetation structure, we consider the presence or

absence of particular bird species and their communities

rather than the rate of development of ecological function. As

hawks and owls prey on mammals, then succession of

mammal species is, in turn, dependent on bird community

dynamics.2

WINTER BIRD COMMUNITIES IN AFFORESTATION: SHOULD WE SPEED UP

OR SLOW DOWN ECOLOGICAL SUCCESSION?

Paul B. Hamel, Daniel J. Twedt, Timothy J. Nuttle,

Christopher A. Woodson, Fred Broerman, and Joseph M. Wahome1

1 Hamel, Research Wildlife Biologist, USDA Forest Service, Southern Research Station, Stoneville, MS; Twedt, Wildlife Biologist, U.S.

Department of the Interior, Geological Survey, Vicksburg, MS; Nuttle, Graduate Research Fellow, Utah State University, Department of Biology

and Ecology Center, Logan, UT; Woodson, Assistant Refuge Manager, U.S. Fish and Wildlife Service, Cruger, MS; Broerman, Wildlife Biologist,

U.S. Fish and Wildlife Service, Bethel, AK; and Wahome, Associate Professor, Mississippi Valley State University, Department of Natural

Sciences & Environmental Health, Itta Bena, MS.

2 Wahome, Joseph. 1999. Unpublished data. On file with: Department of Natural Sciences & Environmental Health, Mississippi Valley State

University, 14000 Highway 82 West, Itta Bena, MS 38941.

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability of

wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.

Abstract—Recent assessments of afforestation on agricultural lands in the Mississippi Alluvial Valley imply the importance

of quickly developing vertical forest structure to benefit wildlife. Examining this assumption, we find that mammals and birds

occur through the full successional sere as targets of proactive management and control. Different species of animals thrive

in structures available at different times during succession. Thus, forest managers’ choices of strategies favor species’

success differentially. Early successional species, particularly those avian communities occurring during winter, have

heretofore been considered only in passing. However, because they occur in areas where herbaceous plants dominate

vegetation structure, these communities include species otherwise rare or absent from the landscape. Extensive afforestation in the Mississippi Alluvial Valley provides ephemeral habitat for birds that winter in herbaceous areas. To provide habitat

for winter birds, managers may wish to consider maintaining large tracts in herbaceous vegetation similar to that occurring 3

to 7 years after cessation of farming activities.
99

A review of literature provides a summary of expected

responses of birds and small mammals to early

successional habitats in MAV landscapes. Empirical

observations compare responses of early successional

wildlife species to four different afforestation treatments. Our

primary data set involves winter bird populations in the first

few years after abandonment of agricultural lands. Nuttle

(1997) investigated bird communities at afforested sites of

different ages in the LMAV that were established by planting

oak (Quercus L. spp.) seedlings. Nuttle (1997) and Nuttle

and Burger (1995) summarize breeding bird communities in

oak plantations, some of which were 30 years old. However,

we have too few data to demonstrate what happens to

species during middle successional conditions occurring 20

to 60 years after initial afforestation.

METHODS

In a comparison of bird occurrence among different methods

of affortestation, we focus on total abundance, density of

individuals, and species richness.

Literature Analysis

The National Council of the Paper Industry for Air and

Stream Improvement, Inc. (NCASI) has published annotated

bibliographies on wildlife, forestry, and habitat in the East

(NCASI 1993, 1999). A keyword search yielded papers

dealing with wildlife response to afforestation, reforestation,

forest restoration, and old-field succession in bottomland

landscapes in the Southeastern United States. However, few

publications treat responses to afforestation in the MAV.

Studies done in other landscapes and regions can

contribute insights as similar species respond to different

conditions.

Empirical Observations of Winter Bird

Communities

Between late October and late April from 1996 to 2000,

Broerman surveyed 8 bird species on 59 afforestation sites

in 9 Mississippi counties. The sites were all former

agricultural land 2 to 9 years after cessation of farming

activities. Some had been planted with tree seedlings,

principally Nuttall oak (Quercus nuttallii Palmer), green ash

(Fraxinus pennsylvanica Marshall), willow oak (Q. phellos

L.), or water oak (Q. nigra L.); some had not. Herbaceous

vegetation on these sites exceeded the height of woody

vegetation. The survey involved using a four-wheeler to flush

birds from herb or grass-dominated vegetation. During this

survey he made 273 trips to the sites. Of these, 271 trips to

57 sites took place in 7 counties.

In 1999 to 2000, Hamel conducted Winter Bird Populations

Studies (WBPS) (Kolb 1965) on the treatment plots in the

Sharkey large-scale restoration experiment near Anguilla,

Sharkey County, MS (Sharkey Site) (Schweitzer and others

1997). Hamel and Woodson made additional observations on

these plots in 1998 and 1999. As a comparison of

afforestation methods, the design of the Sharkey Site was

randomized complete blocks design, involving three

replicates each of three afforestation treatments and a

natural succession and regeneration control (NAT). The

treatments were: (1) sown Nuttall oak acorns (SOW), 2,500

per hectare; (2) planted Nuttall oak seedlings (PLN), 750 per

hectare; and (3) planted cottonwood stem cuttings (NUR),

750 per hectare, followed by underplanting of Nuttall oak

seedlings, 375 per hectare, 2 years later. Individual treatment

plots were approximately 8 ha in extent. Initial establishment

of the treatments was in spring 1995, with the underplanting

in the NUR plots conducted in spring 1997.

Hamel conducted a WBPS on each of the 12 plots on the

Sharkey Site. His WBPS counts were 30 minutes during the

morning hours on 8 days from December 1999 to March

2000. He recorded all birds seen or heard in the vegetation

of the plot, or actively foraging over the plot. He visited 1 of

12 plots in a random sequence. He summarized these

results as average number of individuals encountered per

species per hectare.

We test the hypothesis that early successional species may

not benefit from rapid afforestation as much as from natural

succession. Our null hypothesis for this test is that bird

species dependent on herbaceous vegetation will be found

in the faster growing cottonwood plantations as well as in

other treatments that are accumulating woody structure

more slowly. Our data for this test consist of the abundance

and occurrence of the individual species recorded on the

WBPS in the different treatment plots of the Sharkey Site.

We test the hypothesis that the bird community that

develops in afforestation areas reflects the presence of

separate elements of vegetation structure. We compare the

occurrence of bird species among the treatments on the

Sharkey Site in terms of their foraging preferences from

Hamel (1992) for this test. The null hypothesis for this

examination is that no association between foraging

substrate and occurrence by treatment will be apparent

among the birds.

We test the hypothesis that accumulating vegetation

structure determines the occurrence of birds in afforestation

treatments. Physical vegetation structure accumulates on

the experimental treatments at a predictable sequence,

measured by the height of the woody vegetation,

NUR>PLN>SOW>NAT. Using the data on species richness

and density of individual species among treatments from the

Sharkey Site, we conduct analysis of variance (ANOVA).

Our analysis uses PROC GLM or PROC ANOVA, as a

randomized complete block design with treatment or block

and treatment as main effects (SAS Institute Inc. 1989).

Significance was accepted at p< 0.05; means testing was

conducted using Duncan’s multiple comparison of means.

For tests of the difference of abundance of individual

species among treatments, we accepted significance at p<

0.10 experiment wide with sequential Bonferroni correction.

RESULTS AND DISCUSSION

Literature Analysis

Of 38 papers in the literature search (citations in table 1),

13 deal exclusively with single species response to habitat

restoration, 6 with songbirds in general, 2 with waterfowl,

3 with general vertebrate response to reforestation, 3 with

effects of streamside restoration on aquatic communities,

3 with effects of fire on habitat restoration, and 4 with the
100

importance of reforestation to the maintenance of old-growth

conditions on the landscape. Three papers discuss the role

of animals, specifically beaver (Castor canadensis Kuhl),

white-tailed deer [Odocoileus virginianus (Zimmermann)],

and fruit-eating birds as agents modifying plant community

composition. Predetermined objectives are essential in

restoration planning, as in all other land management, a

point made by a single paper (McCullough 1994). No paper

explicitly treats early secondary succession on abandoned

farmland in the MAV. The greater proportion of the papers

indicate that restoration that proceeds faster from open

ground to closed-canopy forest is more effective than that

which proceeds more slowly.

Twelve of the papers treat bottomland hardwood or other

lowland forest types. Among these, Ribbeck and Hunter

Table 1—Literature review of wildlife responses to afforestation, reforestation, and forest restoration

Category Citation

Single species responses to restoration

Black bear (Ursus americanus Pallas) Weaver and others 1990, Weaver and Pelton 1994

Red-cockaded woodpecker [Picoides borealis (Vieillot)] Cantrell and others 1995, Conner and Rudolph 1995,

Gaines and others 1995, Watson and others 1995,

Wilson and others 1995

American woodcock (Scolopax minor Gmelin) Sepik and Blumenstock 1993

Wild turkey (Meleagris gallopavo L.) Dickson 1992

Florida scrub-jay [Aphelocoma coerulescens (Bosc)] Root and others 1995; Schmalzer 1993, 1994;

Schmalzer and others 1994

Beaver pond reforestation Houston 1991

Deer browse affect on reforestation Anderson and Katz 1993

Songbirds in general Bielefeldt and Rosenfield 1994, Nuttle and Burger 1995,

Ribbeck and Hunter 1994, Tomlinson 1977,

Wesley and others 1976

Restoration for songbirds Ribbeck and Hunter 1994

Reforestation and waterfowl Kaminski and others 1993, Reinecke 1994

Bird dispersal of fruiting plants into restoration areas Robinson and Handel 1993

General vertebrate response to reforestation Askins and Philbrick 1987, Litvaitis 1993

Using vertebrates to assess cumulative impacts Croonquist and Brooks 1991

Land management planning McCullough 1994

Forest restoration of streamside zones,

effect on aquatic species Flebbe and Dolloff 1995, O’Brien-White and Thomason 1995,

Sweeney 1993

Fire effects in restoration Fitzgerald and Tanner 1992, Provencher and others 1995,

Simberloff 1993

Old-growth forest restoration, biodiversity conservation Harris and Scheck 1991; Mladenoff and others 1993, 1994;

Vora 1994

Clearcuts Hurst and Bourland 1996, Mitchell 1989
101

(1994) note that many bird species of highest conservation

priority in the MAV are late successional species; therefore,

rapid afforestation will benefit them.

Wesley and others (1976) studied winter birds in 44

cottonwood plantations aged 4, 5, or 6 years old, within

primarily forested landscapes in the MAV. In an associated

study, Tomlinson (1977) conducted WBPS on five plots,

including three on two cottonwood plantations and two

mature hardwood controls. Six species preferred the

cottonwood plantations, 4 showed no preference, and 15

preferred the natural stands (Tomlinson 1977, Wesley and

others 1976). Wesley and others (1976) compared bird

communities in plantations to those in nearby mature natural

stands, but not in plowed or harvested agricultural fields. The

winter bird community of agricultural fields in the MAV is a

simple one (table 2).

In 1994, Twedt conducted three WBPS in planted

cottonwoods surrounded by agricultural fields on Fitler

Managed Plantation, near Fitler, MS.3 He found 35 species

on the plots, 4 of which were not found by Tomlinson (1977)

or Wesley and others (1976): red-shouldered hawk, northern

bobwhite, chipping sparrow, and Lincoln’s sparrow.

Litvaitis (1993) and Sepik and Blumenstock (1993) note that

as landscapes change from primarily agricultural to primarily

forested settings, species of early successional vegetation

benefit in the short term and ultimately decline to the point

that specific manipulative action is required to maintain their

populations on the landscape. This test does not refute the

hypothesis that early successional species will benefit less

from afforestation than later successional species.

Empirical Observations

Survey—Table 2 lists the winter bird community Twedt

observed in agricultural fields in the MAV. In his surveys of

early successional habitats, Broerman tracked occurrence of

eight species of birds (table 3). These species are relatively

rare in the MAV and of specific conservation interest. Two

species—sedge wren and Le Conte’s sparrow—occurred on

more than half of the surveyed sites (table 3). Each species

is associated with grasslands or with herbaceous vegetation

in the earliest stages of forest succession (Hamel 1992);

neither appears in the cottonwood plantations studied by

Wesley and others (1976), Tomlinson (1977), or Twedt.4

Experimental test-species occurrence—Vegetation

structure on the plots at Sharkey Site differs by treatment at

5 years after establishment. On the NUR treatments,

cottonwood trees approach 10 m or more in height. Nuttall

oak seedlings are approximately 3 to 4 m tall in the PLN and

1 to 3 m tall in the SOW. On the NAT, few woody stems

exceed the 1- to 3-m height of the herbaceous vegetation.

These differences in structure are consistent with the

Table 2—Bird species commonly found in fallow agricultural fields in the Mississippi Delta in the winter from

three Winter Bird Populations Studies conducted in 1994a

Common and scientific names Bird densityb

 Per km2

Great blue heron

Ardea herodias Linn. 0.1 ± 0.1

Northern harrier

Circus cyaneus (L.) .4 ± .2

Cooper’s hawk

Accipiter cooperii (Bonaparte) .4 ± .4

Red-tailed hawk

Buteo jamaicensis (Gmelin) .7 ± .3

American kestrel

Falco sparverius L. .1 ± .1

Killdeer

Charadrius vociferus L. 23.6 ± 21.0

Common snipe

Gallinago gallinago (L.) .1 ± .1

Rock dove

Columba livia Gmelin 9.7 ± 9.7

Mourning dove

Zenaida macroura (L.) 17.1 ± 11.1

Red-bellied woodpecker

Melanerpes carolinus (L.) .2 ± .1

Loggerhead shrike

Lanius ludovicianus L. 1.9 ± 1.3

Blue jay

Cyanocitta cristata (L.) .9 ± .7

Horned lark

Eremophila alpestris (L.) 17.3 ± 12.1

European starling

Sturnus vulgaris L. 189.0 ± 189.0

Vesper sparrow

Pooecetes gramineus (Gmelin) .4 ± .4

Savannah sparrow

Passerculus sandwichensis (Gmelin) 61.9 ± 57.7

Song sparrow

Melospiza melodia (Wilson) 3.8 ± 2.7

White-throated sparrow

Zonotrichia albicollis (Gmelin) 10.2 ± 9.5

White-crowned sparrow

Z. leucophrys (Forster) .7 ± .7

Dark-eyed junco

Junco hyemalis (L.) .1 ± .1

Northern cardinal

Cardinalis cardinalis (L.) 8.2 ± 7.1

Red-winged blackbird

Agelaius phoeniceus (L.) 2.3 ± 1.4

Eastern meadowlark

Sturnella magna (L.) 1.2 ± 1.2

Mean density 350.4 ± 210.9

Mean species richness 13.3 ± 2.8

a Twedt, Daniel. 1994. Unpublished field notes. On file with: U.S.

Department of Interior, Geological Survey, Patuxent Wildlife

Research Center, 2524 South Frontage Road, Vicksburg, MS.

b Plus or minus standard error.

3

 Twedt, Daniel. 1999. Unpublished data. On file with: U.S.

Department of the Interior, Geological Survey, Patuxent Wildlife

Research Center, 2524 South Frontage Road, Vicksburg, MS.

4 Twedt, Daniel. 1994. Unpublished data. On file with: U.S.

Department of the Interior, Geological Survey, Patuxent Wildlife

Research Center, 2524 South Frontage Road, Vicksburg, MS.
102

intensity of the management of the plots at establishment,

with the age of the propagules when planted, and with the

growth rates of the different species planted. Vegetation

structure has accumulated more rapidly in plots in which

more intense effort was made to establish woody vegetation.

Hamel found a total of 51 bird species in the 1999 to 2000

WBPS at the Sharkey Site (table 4). We examine the

occurrence of individual species as a function of their

association with vegetation structure, as well as with respect

to their conservation priority. The ANOVA of species richness

among treatments revealed that significantly more species

(30 plus or minus 4.6 S.D. vs. 11.7 plus or minus 1.8 S.D.)

occurred in the NUR treatment than in the others (F = 54.2,

d.f. = 5,6, P<0.0001 (table 4). The ANOVA of total

abundance was similar among treatments (F = 1.13, d.f. =

5,6, P<0.43 (table 4). Thus 5 years after establishment, the

treatment with the greatest development of vegetation

structure (NUR) harbored the greatest number of bird

species, although it did not harbor a greater number of

individuals than the other treatments. Among the 28 species

found only in the NUR treatment, 12 foraged on trees

(Hamel 1992). Four of the five species never found in the

NUR foraged on the ground or herbaceous vegetation.

These results are consistent with the hypothesis that bird

species occurrence reflects physical vegetation structure.

As vegetation structure develops, avian species appear that

are associated with that structure, such as Eastern phoebe

and yellow-rumped warbler. These birds, which forage from

trees or at variable heights, were recorded only in the NUR

plots in 1998 to 1999. In 1999 to 2000, these birds occurred

rarely in the emerging woody vegetation of the other

treatments, particularly the PLN. In both years, species

associated primarily with open vegetation also occurred

beneath the trees in the NUR, notably song sparrow, swamp

sparrow, and red-winged blackbird.

Other species associated with open habitats, such as

Northern harrier, sedge wren, savannah sparrow, Le Conte’s

sparrow, and eastern meadowlark rarely occurred in the

NUR treatment. For example, sedge wren was found

regularly in all open habitats (54 of 72 visits), but in the NUR

only once (of 24 visits) in a very open spot near the edge

where flooding had killed some of the planted cottonwoods.

An association between occurrence and physical vegetation

structure does not explain why Le Conte’s sparrow occurred

primarily in one spot overlapping the border of a NAT and a

PLN plot. The distribution of this bird does not seem to be a

straightforward response to the afforestation treatments in

this experiment.

Experimental test-species conservation priority—The

Partners in Flight (PIF) offers a generally accepted system

for recording bird species conservation priority (Partners in

Flight 2000). The PIF system gives each species a priority

ranking score based on several aspects of occurrence,

abundance, and threats to its population. These priority

scores are recorded for physiographic areas in which a

species breeds, but not for areas where it only winters

(Colorado Bird Observatory 1999); the maximum possible

score is 35 the minimum is 7 (Carter and others 2000).

When possible, we used the PIF conservation priority scores

from the MAV for this analysis; where scores were

unavailable for the MAV, we used the score from a

representative physiographic region in the breeding range of

the species. Nuttle and others (2000) have used PIF

concern scores similarly.

While bird species richness increases with vegetation

structure in the Sharkey Site, conservation priority of

individual species does not (table 4). Using the PIF priority

rankings, several observations are suggestive. First, the

average conservation priority of all 51 species recorded on

the WBPS is 15.2. The average ranking of all species found

in common among all treatments is a similar 15.5. Second,

28 species that were unique to the NUR treatment average

14.6, a slightly lower value. Third, 5 species not found in the

NUR average 16.6, a value higher than the average. When

the sedge wren is included in this group, average priority

ranking of species found in the other treatments increases to

17.3. Fourth, among the six species of highest conservation

priority, two never occurred in the NUR and one occurred

only in the NUR. Therefore, rapid afforestation provides

winter habitat for a number of species quickly, at the

expense of a few high-priority species found in early

successional habitats.

CONCLUSIONS

Extensive restoration of forests in the MAV may provide

demonstrable, albeit unintended, benefits to birds that winter

within afforested sites in early successional stages. This data

set illustrates that bird species composition in the MAV

Table 3—Bird species observed on 273 trips to 59

afforestation sites in 9 counties in Mississippi from

October 17, 1996 to April 24, 2000, by F. Broerman

Common and scientific names a Sites Birds/trip

 - - - - Number - - - -

American bittern

Botaurus lentiginosus (Rackett) 14 2.3

Yellow rail

Coturnicops noveboracensis

(Gmelin) 4 1.0

Sora

Porzana carolina (L.) 16 5.1

Short-eared owl

Asio flammeus (Pontoppidan) 21 5.6

Sedge wren

Cistothorus platensis (Latham) 33 2.5

Marsh wren

C. palustris (Wilson) 20 2.4

Field sparrow

Spizella pusilla (Wilson) 10 3.1

Le Conte’s sparrow

Ammodramus leconteii (Audubon) 36 4.5

a

 In addition to the species listed, Broerman did not count

individuals of savannah, swamp, and song sparrows, or of eastern

meadowlarks, all of which were numerous on the surveyed sites.
103

Table 4—Foraging site, abundance on different afforestation treatments, and conservation priority of bird species

recorded on four afforestation treatments during winter 1999 to 2000 at the Sharkey Large-Scale Demonstration Project,

Sharkey County, Mississippi

Abundance in treatmentd

Foraging Conservation priority

Common and scientific names sitee NAT SOW PLN NUR ratingf

Northern harrier

Circus cyaneus (L.) G, H 13 13 13 1 21aa

Red-tailed hawk

Buteo jamaicensis (Gmelin) G, H 2.1 9.9 2.6 2.1 12

American kestrel

Falco sparverius L. G .5 0 0 .5 12

Northern bobwhite

Colinus virginianus (L.) G 1 0 0 0 20

Common snipe

Gallinago gallinago (L.) W 2.1 1 .5 0 13aa

American woodcock

Scolopax minor Gmelin G 0 0 .5 .5 19

Mourning dove

Zenaida macroura (L.) G 0 0 0 2.6 14

Great horned owl

Bubo virginianus (Gmelin) G 0 0 0 .5 12

Barred owl

Strix varia Barton V 0 0 0 2.6 16

Red-bellied woodpeckera

Melanerpes carolinus (L.) T 0 0 0 2.1 17

Yellow-bellied sapsuckera

Sphyrapicus varius (L.) T 0 0 0 1 16aa

Downy woodpeckera

Picoides pubescens (L.) T, B 0 0 0 14.6 14

Hairy woodpeckera

P. villosus (L.) T 0 0 0 1.6 14

Northern flickera

Colaptes auratus (L.) G, T 0 0 0 7.8 16

Eastern phoebe

Sayornis phoebe (Latham) V .5 0 2.1 19.3 15

Loggerhead shrike

Lanius ludovicianus L. G 4.7 3.1 8.3 .5 19

Blue jay

Cyanocitta cristata (L.) T 0 0 0 2.1 13

American crow

Corvus brachyrhynchos Brehm G 0 0 0 .5 10

Carolina chickadeea

Poecile carolinensis (Audubon) T 0 0 0 13 20

Tufted titmouseb

Baeolophus bicolor (L.) T 0 0 0 .5 14

Carolina wrenb

Thryothorus ludovicianus (Latham) G, B 0 0 0 9.9 17

Winter wrenb

Troglodytes troglodytes (L.) G 0 .5 0 8.8 14bb

Sedge wren

Cistothorus platensis (Latham) H 23.4 23.7 21.4 .5 21

Golden-crowned kinglet

Regulus satrapa Lichtenstein T 0 0 0 3.1 17bb

continued
104

Table 4—Foraging site, abundance on different afforestation treatments, and conservation priority of bird species

recorded on four afforestation treatments during winter 1999 to 2000 at the Sharkey Large-Scale Demonstration Project,

Sharkey County, Mississippi (continued)

Abundance in treatmentd

Foraging Conservation priority

Common and scientific names sitee NAT SOW PLN NUR ratingf

Ruby-crowned kingleta

Regulus calendula (L.) T, B 0 0 0 7.3 16bb

Eastern bluebirda

Sialia sialis (L.) G 0 0 0 2.1 14

Hermit thrusha

Catharus guttatus (Pallas) G, B 0 0 0 3.1 16bb

American robina

Turdus migratorius G, B, T 0 0 0 13 9

Northern mockingbird

Mimus polyglottos (L.) G, B 0 0 .5 0 14

Brown thrasher

Toxostoma rufum (L.) G, B 0 0 0 .5 17

Cedar waxwing

Bombycilla cedrorum Vieillot B, T 0 0 0 2.1 12aa

Yellow-rumped warblera

Dendroica coronata (L.) V, T 0 0 .5 65.6 16bb

Palm warbler

Dendroica palmarum (Gmelin) G, H, B 0 0 0 .5 16cc

Common yellowthroat

Geothlypis trichas (L.) H, B .5 2.6 1 .5 16

Eastern towhee

Pipilo erythrophthalmus (L.) G, B 0 0 0 .5 15

Field sparrow

Spizella pusilla (Wilson) H 0 2.1 0 .5 20

Savannah sparrow

Passerculus sandwichensis (Gmelin) G, H 94.8 74.5 140.1 0 13aa

Le Conte’s sparrow

Ammodramus leconteii (Audubon) H 1.6 0 4.7 0 23aa

Fox sparrow

Passerella iliaca (Merrem) G 0 .5 .5 15.1 12bb

Song sparrowc

Melospiza melodia (Wilson) G, H 68.2 78.1 53.1 30.2 12

Swamp sparrowc

M. georgiana (Latham) H 78.6 86.4 124 55.2 16aa

White-throated sparrowc

Zonotrichia albicollis (Gmelin) G, B 0 0 0 43.2 14cc

White-crowned sparrow

Z. leucophrys (Forster) G, H 0 0 0 2.1 12bb

Dark-eyed juncoc

Junco hyemalis (L.) G 0 0 0 10.4 14cc

Northern cardinal

Cardinalis cardinalis (L.) G, B 0 0 0 5.2 12

Red-winged blackbirdc

Agelaius phoeniceus (L.) G, H 241.7 18.2 63.5 252.6 12

Eastern meadowlark

Sturnella magna (L.) G, H 50 47.9 61.5 4.2 17

continued
105

Table 4—Foraging site, abundance on different afforestation treatments, and conservation priority of bird species

recorded on four afforestation treatments during winter 1999 to 2000 at the Sharkey Large-Scale Demonstration Project,

Sharkey County, Mississippi (continued)

Abundance in treatmentd

Foraging Conservation priority

Common and scientific names sitee NAT SOW PLN NUR ratingf

Rusty blackbird

Euphagus carolinus (Müller) G 0 0 0 8.3 16cc

Brewer’s blackbird

E. cyanocephalus (Wagler) G 0 0 0 1.6 15aa

Common grackle

Quiscalus quiscula (L.) G 0 0 .5 222.4 16

American goldfinchc

Carduelis tristis (L.) V 3.6 0 3.1 16.7 12

Unknown species NA 6.3 5.2 5.2 7.3 NA

Mean total density per treatment NA 592.6 366.7 507.2 865.3 NA

Mean no. of species per treatment NA 11 11 13 30 NA

Total species per treatment NA 17 14 19 47 51

Unique species per treatment NA 1 0 1 28 NA

NA = not applicable; NAT = natural succession and regeneration control; SOW = sown Nuttall oak acorns, 2500 per hectare; PLN = planted

Nuttall oak seedlings, 750 per hectare; NUR = planted cottonwood stem cuttings, 750 per hectare, followed by underplanting of Nuttall oak

seedlings, 375 per hectare, 2 years later.

a

 Species associated with cottonwood plantations, which are listed by Wesley and others (1976) as preferring natural stands.

b

 Species associated with cottonwood plantations, which are listed by Wesley and others (1976) as preferring neither plantations nor natural

stands.

c

 Species listed by Wesley and others (1976) as preferring cottonwood plantations.

d Entries reflect mean abundance as birds per km2

. Boldfaced numbers indicate mean abundance different from other treatments by one-way

analysis of variance with treatment as main effect, adjusted for an experiment-wide error rate of 0.10 using sequential Bonferroni correction.

Standard error values available on request from Hamel (lead author of publication).

e Foraging sites listed as B = bush, G = ground, H = herbaceous vegetation, T = tree, V = various heights, W = water (Hamel 1992).

f

 Conservation priority ratings are the Partners in Flight Concern Scores (Colorado Bird Observatory 1999) for the Mississippi Alluvial Valley, with

the exception of aa = Drift Prairie, bb = Central Rocky Mountains, and cc = Great Lakes Transition physiographic regions.

follows vegetation structure and development rate in

afforestation of abandoned agricultural lands. However, as

woody vegetation develops, some bird species of herbaceous

vegetation disappear. Perhaps more importantly, the early

successional avian species that specialize on herbaceous

vegetation are of higher than average conservation priority

among the birds found in afforestation areas.

Different species of mammals and birds respond positively to

the structure available at different times during succession.

Thus, managers must decide on the species and

communities they wish to favor. Winter avian communities of

early successional stages have been considered heretofore

only in passing. However, because birds occur in areas

where vegetation structure is dominated by nonagricultural

herbaceous plants, they include species otherwise rare or

absent from the MAV landscape. To provide bird habitat

through the full successional sere, proactive managers may

wish to dedicate certain areas to natural succession or to

maintenance of herbaceous vegetation.

Forest land managers and biodiversity preservationists

make clear the essential relationship between management

practices and management objectives (McCullough 1994,

Noss 1999). Management (1) is a conscious, goal-directed

activity with goals specified in advance, and (2) employs

practices to achieve goals within an acceptable period of

time. Managers can apply a variety of afforestation methods

under different conditions to achieve different management

objectives. Managerial discretion influences the rate of

structural development of afforestation efforts. Inclusion of

management goals to produce habitats in afforestation for

early successional species is certainly possible, as is

inclusion of management goals to maximize the rate of

development of forest vegetation structure and the birds

associated with it.

ACKNOWLEDGMENTS

This work has benefited from the generous field assistance

of the staff and residents of Sunflower Landing, Dublin, MS.

Mark Woodrey and Jeff Buler of the Mississippi Museum of
106

Natural Science aided in the fieldwork, as did a number of

other collaborators on the Sharkey Large-Scale Restoration

Project.

LITERATURE CITED

Anderson, R.C.; Katz, A.J. 1993. Recovery of browse-sensitive tree

species following release from white-tailed deer Odocoileus

virginianus Zimmerman browsing pressure. Biological Conservation. 63(3): 203–208.

Askins, R.A.; Philbrick, M.J. 1987. Effect of changes in regional

forest abundance on the decline and recovery of a forest bird

community. Wilson Bulletin. 99(1): 7–21.

Bauch, J.; Dunisch, O.; Gasparotto, L., eds. 1999. Investigations

on tree species suitable for the recultivation of degraded land

areas in central Amazonia: Brazilian-German cooperation within

the project ‘studies on human impacts on forests and floodplains in the tropics’ (SHIFT) supported by the Ministry of

Education, Science, Research, and Technology, Bonn, FRG

(0339638) and the Conselho Nacional de Desenvolvimento

Cientifico e Tecnologica (CNPq), Brasilia, Brazil (ENV 42) and

Instituto Brasileiro do Meio Ambiente e Recursos Naturais

Renovaveis (IBAMA). No. 193. Hamburg, Germany:

Mitteilungen der Bundesforschungsanstalt fur Forst- und

Holzwirtschaft. 138 p.

Bielefeldt, J.; Rosenfield, R.N. 1994. Summer birds of conifer

plantations in southeastern Wisconsin. Passenger Pigeon. 56(2):

123–135.

Boyle, J.R. 1999. Planted forests: views and viewpoints. New

Forests. 17(1/3): 5–9.

Burgess, P.J. 1999. Effects of agroforestry on farm biodiversity in the

UK. Scottish Forestry. 53(1): 24–27.

Cannell, M.G.R. 1999a. Environmental impacts of forest monocultures: water use, acidification, wildlife conservation, and carbon

storage. New Forests. 17(1/3): 239–262.

Cannell, M.G.R. 1999b. Forests, Kyoto and climate. Outlook on

agriculture. 28(3): 171–177.

Cantrell, M.A.; Britcher, J.J.; Hoffman, E.L. 1995. Red-cockaded

woodpecker management initiatives at Fort Bragg Military

Installation. In: Kulhavy, D.L.; Hooper, R.G.; Costa, R., eds. Redcockaded woodpecker: recovery, ecology and management.

Red-cockaded woodpecker symposium III: species recovery,

ecology and management; 1993 January 24–28; North

Charleston, SC. Nacogdoches, TX: Stephen F. Austin State

University, Center for Applied Studies, College of Forestry: 89–

97.

Carter, M.F.; Hunter, W.C.; Pashley, D.N.; Rosenberg, K.V. 2000.

Setting conservation priorities for landbirds in the United States:

the partners in flight approach. Auk. 117(2): 541–548.

Chang, ChingCheng. 1999. Carbon sequestration cost by afforestation in Taiwan. Environmental Economics and Policy Studies.

2(3): 199–213.

Colorado Bird Observatory. 2000. Colorado Bird Observatory PIF

database. Barr Lake State Park, CO: Colorado Bird Observatory.

[Available on the Web: http://www.cbobirds.org/pif/index.html].

Conner, R.N.; Rudolph, D.C. 1995. Losses of red-cockaded

woodpecker cavity trees to southern pine beetles. Wilson

Bulletin. 107(1): 81–92.

Croonquist, M.J.; Brooks, R.P. 1991. Use of avian and mammalian

guilds as indicators of cumulative impacts in riparian-wetland

areas. Environmental Management. 15(5): 701–714.

DeGraaf, R.M. 1987. Managing northern hardwoods for breeding

birds. In: Nyland, R.D., ed. Proceedings of the silvicultural

symposium managing northern hardwoods; 1986 June 23–25;

Syracuse, NY. Misc. Publ. 13 (ESF 87–002). Syracuse, NY:

Syracuse University, Faculty of Forestry: 348–362.

DeGraaf, R.M.; Yamasaki, M.; Leak, W.B.; Lanier, J.B. 1992. New

England wildlife: management of forested habitats. Gen. Tech.

Rep. NE–144. Radnor, PA: U.S. Department of Agriculture,

Forest Service, Northeastern Forest Experiment Station. 271 p.

Dickson, J.G., ed. 1992. The wild turkey: biology and management.

Harrisburg, PA: Stackpole Books. 463 p.

Fitzgerald, S.M.; Tanner, G.W. 1992. Avian community response to

fire and mechanical shrub control in south Florida. Journal of

Range Management. 45(4): 396–400.

Flebbe, P.A.; Dolloff, C.A. 1995. Trout use of woody debris and

habitat in Appalachian wilderness streams of North Carolina.

North American Journal of Fisheries Management. 15: 579–590.

Gaines, G.D.; Franzreb, K.E.; Allen, D.H. [and others]. 1995. Redcockaded woodpecker management on the Savannah River Site:

a management/research success story. In: Kulhavy, D.L.;

Hooper, R.G.; Costa, R., eds. Red-cockaded woodpecker:

recovery, ecology and management. Red-cockaded woodpecker

symposium III: species recovery, ecology and management;

1993 January 24–28; North Charleston, SC. Nacogdoches, TX:

Stephen F. Austin State University, Center for Applied Studies,

College of Forestry: 81–88.

Hamel, P.B. 1992. A land manager’s guide to the birds of the South.

Atlanta: U.S. Department of Agriculture, Forest Service,

Southern Region; Chapel Hill, NC: The Nature Conservancy,

Southeastern Region. 367 p.

Harrington, C.A. 1999. Forests planted for ecosystem restoration or

conservation. New Forests. 17(1/3): 175–190.

Harris, L.D.; Scheck, J. 1991. From implications to applications: the

dispersal corridor principle applied to the conservation of

biological diversity. In: Saunders, D.A.; Hobbs, R.J., eds. Nature

conservation 2: the role of corridors. Chipping Norton, Australia:

Surrey Beatty & Sons: 189–220.

Helmer, W. 1999. [Establishment of broadleaved floodplain forest].

Opkomst van het hardhoutooibos. Nederlands

Bosbouwtijdschrift. 71(3): 130–134.

Houston, A.E. 1991. Beaver control and reforestation of drained

beaver impoundments. Knoxville, TN: University of Tennessee.

143 p. Ph.D. dissertation.

Hurst, G.A.; Bourland, T.R. 1996. Breeding birds on bottomland

hardwood regeneration areas on the Delta National Forest.

Journal of Field Ornithology. 67: 181–187.

James, F.C. 1971. Ordinations of habitat relationships among

breeding birds. Wilson Bulletin. 83: 215–236.

Joslin, J.D.; Schoenholtz, S.H. 1997. Measuring the environmental

effects of converting cropland to short-rotation woody crops: a

research approach. Biomass and Bioenergy. 13(4/5): 301–311.

Kaminski, R.M.; Alexander, R.W.; Leopold, B.D. 1993. Wood duck

and mallard winter microhabitats in Mississippi hardwood

bottomlands. Journal of Wildlife Management. 57(3): 562–570.
107

Kolb, H. 1965. The Audubon winter bird population study. Audubon

Field Notes. 19(3): 432–434.

Litvaitis, J.A. 1993. Response of early successional vertebrates to

historic changes in land use. Conservation Biology. 7(4): 866–

873.

McCullough, D.R. 1994. Importance of population data in forest

management planning. Forestry Chronicle. 70(5): 533–537.

Mitchell, L.J. 1989. Effects of clearcutting and reforestation on

breeding bird communities of baldcypress-tupelo wetlands.

Raleigh, NC: North Carolina State University. 92 p. M.S. thesis.

Mladenoff, D.J.; White, M.A.; Crow, T.R.; Pastor, J. 1994. Applying

principles of landscape design and management to integrate oldgrowth forest enhancement and commodity use. Conservation

Biology. 8(3): 752–762.

Mladenoff, D.J.; White, M.A.; Pastor, J.; Crow, T.R. 1993. Comparing

spatial pattern in unaltered old-growth and disturbed forest

landscapes. Ecological Applications. 3(2): 294–306.

Mueller, A.J.; Loesch, C.R.; Twedt, D.J. 1995. Development of

management objectives for breeding birds in the Mississippi

Alluvial Valley. In: Bonney, R.; Pashley, D.N.; Cooper, R.J.; Niles,

L., eds. Proceedings of the Partners in Flight international

workshop strategies for bird conservation: the Partners in Flight

planning process; 1995 October 1–5; Cape May, NJ. Ithaca, NY:

Cornell Lab of Ornithology. [Various pagination]. [Available on the

Web: http://www.birds.cornell.edu/pifcapemay/mueller.htm].

National Council of the Paper Industry for Air and Stream Improvement, Inc. (NCASI). 1993. Forestry, wildlife, and habitat in the

East: an annotated bibliography, 1986–1990. NCASI Tech. Bull.

651. Research Triangle Park, NC: National Council of the Paper

Industry for Air and Stream Improvement, Inc. 340 p.

National Council of the Paper Industry for Air and Stream Improvement, Inc. (NCASI). 1999. Forestry, wildlife, and habitat in the

East: an annotated bibliography, 1991–1995. NCASI Tech. Bull.

781. Research Triangle Park, NC: National Council of the Paper

Industry for Air and Stream Improvement, Inc. 753 p.

Niyaz, S.; Nigi, T.; Koshika, K. 1999. [Growth of planted subsidized

trees in natural forests in the central Tianshan Mountains and

damage caused by grazing cattle]. Research Bulletin of the

Hokkaido University Forests. 56(1): 94–104.

Noss, R.F. 1999. Assessing and monitoring forest biodiversity: a

suggested framework and indicators. Forest Ecology and

Management. 115(2/3): 135–146.

Nuttle, T.J. 1997. Response of breeding bird communities to

afforestation of hardwood bottomland sites in Mississippi.

Mississippi State, MS: Mississippi State University. 68 p. M.S.

thesis.

Nuttle, T.J.; Burger, L.W. 1995. Breeding bird response to hardwood

bottomland restoration in the Mississippi Alluvial Valley.

[Abstract]. In: Program of Wildlife Society second annual

conference: excellence in wildlife stewardship through science

and education; 1995 September 12–17; Portland, OR: 92–93.

Nuttle, T.; Leidolf, A.; Burger, L.W. 2000. Using Partners in Flight

species concern scores to compare the conservation value of

bird communities [Abstract]. In: One hundred eighteenth stated

meeting of the American Ornithologists’ Union; 2000 August 14–

19; St. John’s, Newfoundland, Canada. St. John’s, Newfoundland, Canada: American Ornithologists’ Union: 26.

O’Brien-White, S.; Thomason, C.S. 1995. Evaluating fish habitat in a

South Carolina watershed using GIS. Proceedings of the annual

conference of the Southeastern Association of Fish and Wildlife

Agencies; 1995 September 23–27; Nashville, TN. Tallahassee,

FL: Southeastern Association of Fish and Wildlife Agencies. 49:

153–166.

Pande, V.C.; Nambiar, K.T.N.; Singh, H.B. 1999. Afforestation on

community land in the semi-arid tropics of Gujarat—an economic analysis. Indian Forester. 125(2): 212–218.

Partners in Flight. 2000. Partners in Flight – U.S. home page.

Patuxent National Wildlife Refuge, MD: U.S. Department of the

Interior, Biological Resources Division. [Various pagination].

[Available on the Web: http://www.partnersinflight.org].

Provencher, L.; Demarest, D.; Brennan, L. [and others]. 1995. Avian

community responses to herbaceous invertebrates and

vegetation in burned old-growth and fire-suppressed longleaf

pine habitats: experimental design and pretreatment results

[Abstract]. In: Program of Wildlife Society second annual

conference: excellence in wildlife stewardship through science

and education; 1995 September 12–17; Portland, OR. Portland,

OR: The Wildlife Society: 105.

Reinecke, K.J. 1994. Regional waterfowl habitat trends and

implications for neotropical migratory birds. In: Smith, W.P.;

Pashley, D.N., comps., eds. A workshop to resolve conflicts in

the conservation of migratory landbirds in bottomland hardwood

forests; 1993 August 9–10; Tallulah, LA. Gen. Tech. Rep. SO–

114. New Orleans: U.S. Department of Agriculture, Forest

Service, Southern Forest Experiment Station: 20–24.

Ribbeck, K.F.; Hunter, W.C. 1994. Is reforestation an adequate

restoration of bottomland hardwood functions for the needs of

neotropical migratory birds? In: Smith, W.P.; Pashley, D.N.,

comps., eds. A workshop to resolve conflicts in the conservation

of migratory landbirds in bottomland hardwood forests; 1993

August 9–10; Tallulah, LA. Gen. Tech. Rep. SO–114. New

Orleans: U.S. Department of Agriculture, Forest Service,

Southern Forest Experiment Station: 25–26.

Robinson, G.R.; Handel, S.N. 1993. Forest restoration on a closed

landfill: rapid addition of new species by bird dispersal. Conservation Biology. 7(2): 271–278.

Root, K.V.; Breininger, D.R.; Bowman, R.; Swain, H.M. 1995. What

can population viability analysis (PVA) tell us about the probable

success of a spatially explicit reserve design: the Florida scrub

jay in Brevard County as a case study [Abstract]. Bulletin of the

Ecological Society of America. 76(2 Suppl. Part 2): 229.

SAS Institute Inc. 1989. SAS/STAT user’s guide. Version 6. 4th ed.

Cary, NC: SAS Institute Inc. 846 p. Vol. 2.

Schmalzer, P.A. 1993. Characteristics of long unburned stands of

oak-saw palmetto scrub [Abstract]. Bulletin of the Ecological

Society of America. 74(2 Suppl.): 428.

Schmalzer, P.A. 1994. Restoring unburned shrublands: an example

using cutting and burning in oak-saw palmetto scrub [Abstract].

Bulletin of the Ecological Society of America. 75(2 Suppl.): 206.

Schmalzer, P.A.; Breininger, D.R.; Adrian, F.W. [and others]. 1994.

Development and implementation of a scrub habitat compensation plan for Kennedy Space Center. Tech. Memo. 109202.

Kennedy Space Center, FL: National Aeronautics and Space

Administration. 54 p.

Scholtens, L.J.R. 1998. Environmental, developmental and financial

risks of tropical timber plantation investment funds. Natural

Resources Forum. 22(4): 271–277.
108

Schweitzer, C.J.; Stanturf, J.A.; Shepard, J.P. [and others]. 1997.

Large-scale comparison of reforestation techniques commonly

used in the Lower Mississippi Alluvial Valley: first year results. In:

Pallardy, S.G.; Cecich, R.A.; Garrett, H.; Johnson, P.S., eds.

Proceedings of the 11th central hardwood forest conference; 1997

March 23–26; Columbia, MO. Gen. Tech. Rep. NC–188. St. Paul,

MN: U.S. Department of Agriculture, Forest Service, North

Central Forest Experiment Station: 313–320.

Sepik, G.F.; Blumenstock, B. 1993. Management and research on the

American woodcock at the Moosehorn National Wildlife Refuge

[Abstract]. In: Longcore, J.R.; Sepik, G.F., eds. Proceedings of the

eighth American woodcock symposium; 1993; West Lafayette, IN.

Biol. Rep. 16. Washington, DC: U.S. Department of the Interior,

Fish and Wildlife Service, National Biological Survey: 136–137.

Simberloff, D. 1993. Effects of fragmentation on some Florida

ecosystems, and how to redress them. In: Saunders, D.A.;

Hobbs, R.J.; Ehrlich, P.R., eds. Nature conservation 3: reconstruction of fragmented ecosystems. Chipping Norton, Australia:

Surrey Beatty & Sons: 179–187.

Sweeney, B.W. 1993. Effects of streamside vegetation on

macroinvertebrate communities of White Clay Creek in Eastern

North America. Proceedings of the Academy of Natural Sciences

of Philadelphia. 144: 291–340.

Vora, R.S. 1994. Integrating old-growth forest into managed

landscapes: a northern Great Lakes perspective. Natural Areas

Journal. 14(2): 113–123.

Watson, J.C.; Hooper, R.G.; Carlson, D.L. [and others]. 1995.

Restoration of the red-cockaded woodpecker population on the

Francis Marion National Forest: three years post Hugo. In:

Kulhavy, D.L.; Hooper, R.G.; Costa, R., eds. Red-cockaded

woodpecker: recovery, ecology and management. Red-cockaded

woodpecker symposium III: species recovery, ecology and

management; 1993 January 24–28; North Charleston, SC.

Nacogdoches, TX: Stephen F. Austin State University, Center for

Applied Studies, College of Forestry: 172–182.

Weaver, K.M.; Pelton, M.R. 1994. Denning ecology of black bears in

the Tensas River basin of Louisiana. International Conference on

Bear Research and Management. 9(1): 427–433.

Weaver, K.M.; Tabberer, D.K.; Moore, L.U., Jr. [and others]. 1990.

Bottomland hardwood forest management for black bears in

Louisiana. Proceedings of the annual conference of the

Southeastern Association of Fish and Wildlife Agencies; 1990

October 21–24; Richmond, VA. Tallahassee, FL: Southeastern

Association of Fish and Wildlife Agencies. 44: 342–350.

Wesley, D.E.; Perkins, C.J.; Sullivan, D.A. 1976. Preliminary

observations of cottonwood plantations as wildlife habitat. In:

Thielges, B.A.; Land, S.B., Jr. eds. Proceedings of the symposium on eastern cottonwood and related species; 1976 September 28–October 2; Greenville, MS. New Orleans: U.S. Department of Agriculture, Forest Service, Southern Forest Experiment

Station: 460–476.

Willoughby, I.; McDonald, H.G. 1999. Vegetation management in

farm forestry: a comparison of alternative methods of inter-row

management. Forestry (Oxford). 72(2): 109–121.

Wilson, C.W.; Masters, R.E.; Bukenhofer, G.A. 1995. Breeding bird

response to pine-grassland community restoration for redcockaded woodpeckers. Journal of Wildlife Management. 59(1):

56–67.
109

INTRODUCTION

Global Learning and Observations to Benefit the

Environment (GLOBE) is an international science and

education program for K through 12 students. Initiated in

1994, GLOBE is currently supported by 87 countries and

over 100 regional support groups, called GLOBE franchises,

across the United States. Scientists involved in the program

develop detailed guidelines for the collection of

environmental data, monitor data quality, use the data in

their own research, and mentor students. The GLOBE

students, under the guidance of trained GLOBE teachers,

collect data using the science protocols, input the data into a

central database via the Internet, and use the data for their

own research.

The dual role of GLOBE as a science and education

program provides a unique educational advantage for

students. The data students collect are of a quality usable by

earth scientists and are easily accessible to anyone via the

World Wide Web. Dr. Roger Bales, Principal Investigator for

GLOBE Hydrology at the University of Arizona, has said,

“GLOBE hydrology measurements fill a critical gap in water

monitoring and assessment efforts in the United States and

worldwide. GLOBE schools sample many smaller streams

and lakes that are underrepresented in the professional

monitoring programs run by government agencies. We have

now reached the point where GLOBE constitutes one of the

largest water-quality networks in the United States and is

certainly the one with the most readily available data”

(GLOBE offline, in press).

The GLOBE scientists along with other scientists who have

already discovered and are using this unique database have

a real interest in encouraging and mentoring these students.

The students are provided with motivation to participate in

science when they know the data they collect are ‘real’ and

being used—not just an assignment to be discarded after

grading. The value of what the students are doing has been

illustrated by the number of scientists who are not directly

involved in GLOBE, but are using GLOBE data or who have

approached GLOBE for help in collecting special

measurements.

David Verbyla of the University of Alaska, Fairbanks, asked

GLOBE students to provide information on budburst in their

area States.“ There is no network of on-the-ground plant

phenology observations to validate growing season models

and estimates derived from satellite data. GLOBE schools’

plant phenology observations will be the only source of

wide-spread growing season observations for research to

better understand climate change” (GLOBE offline, in press).

Students are further motivated by the opportunity to provide

valuable input into their own communities through their own

research. Visualization tools provided on the GLOBE Web

site allow students to graph and map GLOBE data in real

time. New educational materials guide students through the

process of research. The techniques of data collection and

assessment, such as the creation of classified and assessed

land-cover maps from Landsat images, are often the most

current and accurate tools available for assessment of these

regions.

EARTH AS A SYSTEM

GLOBE data students are collecting fall under a variety of

environmental fields including atmosphere, soil, land cover,

and hydrology. An investigation team or teams that include

science and education co-principal investigators lead each

of these areas. The GLOBE Hydrology Investigation Team

includes three scientists from the University of Arizona: Dr.

Roger Bales, Dr. Martha Conklin, and Dr. Katrina Mangin,

and an education co-pi, Cyndy Henzel. The teams work

together to ensure both the scientific integrity of the data

and the educational quality of the learning materials.

GLOBE, STUDENT INQUIRY, AND LEARNING COMMUNITIES

C.L. Henzel1

1

 Education Development, Scientists Organized to Integrate Learning, Inc., 4920 S. Browning Lane, Tucson, AZ 85746.

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability of

wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.

Abstract—The Global Learning and Observations to Benefit the Environment (GLOBE) database is a web-based archive of

environmental data gathered by K through12 students in over 85 countries. The data are gathered under protocols developed by research scientists specializing in various fields of earth science. Students gather information, then enter and

visualize the data via the Internet. GLOBE’s potential is to provide two major components that affect sustainability and

conservation: quality earth system data and science education opportunities for students. First, GLOBE maintains a quality,

accessible database of millions of environmental data at a geographic scale never before attempted. Second, GLOBE is a

real opportunity for students to learn field study and scientific research methods. GLOBE provides easy access to data and

visualization tools via the Internet to help teachers and students understand, interpret, and ask relevant questions about the

earth system. Students who have a greater understanding of how the environmental system in their community operates

can make more informed decisions on sustainability and conservation issues on a larger scale.
110

Students are encouraged to make their observations in the

context of the earth as a system. Rather than only collecting

point data for a water site, students are asked to also collect

and analyze soil samples, record daily atmosphere

measurements, and create detailed land cover maps. The

water data they collect can then be analyzed in the context

of their specific site and other inputs to their system.

Scientists are finding this wide spatial coverage of integrated

data collection an invaluable tool for filling in gaps for earth

system research. Students gain a greater appreciation of

how the intricate involvement of many systems plays a role

in understanding an environment. For instance, how does

the character of your soil and type of land cover help explain

the amount of input and quality of water at your hydrology

site? What happens if the land cover or pattern of

precipitation changes?

GLOBE recognized early in the development of the program

that the exploration of watersheds was an excellent tool to

use to gain a greater understanding of the interplay of the

earth system on a local, regional, or even global scale.

Students were encouraged to identify their local watershed,

supply contextual information about their hydrology

measurements, and interpret their findings in light of other

measurements and environmental observations. Specific

educational activities to achieve these objectives were

difficult to develop in the first years of GLOBE, however, due

to the lack of long-term data sets to use. But, by the

beginning of this year, the GLOBE student archive contained

over 4 million data points.

STUDENT INQUIRY

In 1999, GLOBE educators began development of materials

to help guide teachers and students in the use of GLOBE

data in student research. Through a number of research

seminars for teachers, it was found that although most

science teachers taught the steps of the scientific method,

few felt comfortable leading students to develop appropriate

research questions, collect and analyze data, and draw

conclusions. The visualization tools provided on the GLOBE

Web site were being underutilized: students were collecting

and inputting data but never looking at it or using it for their

own research.

Development of guiding materials to help students do their

own research easily met the dual science and education

roles within GLOBE. Scientists recognized that students

exploring the database and using the data in their own

research were more likely to recognize the value of quality

data and consistency in data collection. Educators in

schools across the country and around the world were

scrambling to meet new standards that included original

science research or inquiry.

Dr. Leon Lederman, Nobel Laureate: “GLOBE is the

quintessentially ideal program for involving kids in science.

GLOBE teaches science content and also the process of

science. Facts are important, but the younger students are,

the more important it is to learn the process of science.

Science isn’t about providing answers as much as it is about

asking questions. As a hands-on program, GLOBE provides

opportunities for teachers and scientists to talk informally with

kids and get them to ask questions” (GLOBE offline, in press).

Data collected by GLOBE students in the area of hydrology

include measurements of transparency, water temperature,

pH, conductivity or salinity, dissolved oxygen, alkalinity, and

nitrate. New measurements for freshwater and saltwater

macroinvertebrates are currently being developed and

should be implemented by the end of 2000. These

measurements, combined with those from the other GLOBE

investigation areas, provide a comprehensive set of data

from which students can develop research questions and

projects. GLOBE, recognizing that any dataset is more

valuable if it can be combined with other datasets, has

provided reference datasets on the GLOBE Web site from

other sources that allow students easy access. In addition,

GLOBE data can easily be pulled out in table format to

export into a spreadsheet or other analysis tool such as

Geographic Information Systems (GIS).

GLOBE-A-THONS AND LONG-TERM MONITORING

The power of GLOBE student inquiry will hopefully be seen

in two areas. First, students are being encouraged to collect

quality, consistent environmental research data. GLOBE was

developed with the idea of creating a large, accessible

environmental data archive that would have both a wide

spatial scale and long-term monitoring. We are also now

encouraging GLOBE-a-thons, or intensive cooperative data

collection by GLOBE students and perhaps other

environmental organizations within a region. These may

provide a snapshot of an entire watershed that may be then

revisited over time.

Second, GLOBE student inquiry is filling a need within the

schools for original student research. New materials will help

guide students through the research process. The GLOBE

Web site is constantly being improved with the addition of

new tools for student visualization and research, more

student data, and more access to other datasets. In addition,

the Web site now provides space for publication of student

investigations, communication being recognized as an

essential component of the research process. Perhaps most

important of all, GLOBE can help influence those students to

become the mentoring GLOBE scientists for future

generations.

THE GLOBE COMMUNITY

In 1995 when GLOBE trained the first teachers in the

program, it envisioned having a GLOBE teacher in every

school. As the program has grown, so has the vision.

Because of the broad, rich content of the program and broad

applications at all age levels, GLOBE began to encourage

the training of teams of teachers in schools. These schools

quickly found the value of local community resources such

as parent volunteers, professional volunteers from other

environmental groups and Federal, State, and local

government agencies. Some of these groups, including such

organizations as universities, museums, and parks,

developed more formal ties to GLOBE and have become

GLOBE franchises. The franchises take on the responsibility

of training teachers and providing follow-up to local schools.

Franchises have also taken the initiative in taking GLOBE

beyond what can be provided at the Federal level.

Franchises in Alabama and Montana have hosted GIS

conferences for GLOBE franchises and trainers. Franchises
111

in Arkansas, South Carolina, Idaho, and Arizona have

worked together to promote the second International

GLOBE Student Conference to be held in Arkansas in June

of 2000. Internationally, GLOBE countries have also been

active in creating the community. For example, Finland

hosted the first International student conference in 1998.

Croatia recently hosted a land-cover symposium.

What this all means to education and the sustainability of

the important data collection being done by the students is

that the students are now part of a vast learning community.

Students, teachers, parents, governments, universities, and

others are all working toward a common goal. Students are

not just on the receiving end with facts to learn and

homework to turn in. They are an important part of a local

effort to understand their own backyards as well as a global

effort to understand how the jigsaw of local regions fits

together.

CONCLUSION

The study of water has long been used as a key to

developing student interest and skills in science. Many

programs regionally, nationally, and internationally, have

provided and still provide quality materials to help students

gain an understanding of key science concepts like the

water cycle and the measurement of water quality. Others

have concentrated their efforts on the importance of

developing an appreciation for the environment and

conservation of resources. Although GLOBE obviously has

some overlapping objectives with many other environmental

programs, the strength of the program is not realized by

competing for student time and resources.

GLOBE teaches students to ask questions based on

observation, gather data in a scientifically rigorous manner,

analyze data, and draw conclusions based on those results,

and communicate the results to others in the community.

This is the way to make good decisions about the

environment. Good decisions will lead to sustainable

development.

REFERENCES

GLOBE offline. [In press]. Scientists are talking about GLOBE;

spring 2000. Washington, DC: The GLOBE Program.
112

USING KEPONE TO EXEMPLIFY THE IMPORTANCE OF NATURAL

VARIABILITY IN ESTIMATING EXPOSURE TO TOXIC CHEMICALS FROM

AQUATIC ENVIRONMENTS

Robert J. Huggett1

1

 Vice President, Office for Research and Graduate Studies, Michigan State University, 232 Administration Building, East Lansing, MI 48824-1046.

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability of

wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.

Abstract—Kepone, decachlorooctahydro-l, 3, 4-metheno-2H-cyclobuta (cd) pentalen-2-one, is a known mammalian

carcinogen. From at least 1967 to 1975 when production stopped, it contaminated the Chesapeake Bay. Action levels for

kepone in seafood were established by the U.S. Environmental Protection Agency, and various species of finfish, oysters,

Crassostrea virginica, and crabs, Callinectes sapidus, often were found to exceed those levels. Detailed sampling and

analyses of biota showed that interspecies variability in concentrations often exceeded an order of magnitude. Further

examination of the data showed that much of the variability could be explained by factors such as sex, spawning cycle, and

migratory patterns. Estimates of human exposure to kepone-contaminated seafood, and, hence, estimates of risk from

consuming it, were quite inaccurate unless natural variability was considered. On the positive side, an understanding of the

factors controlling natural variability provided alternative risk-management options to minimize risk by decreasing exposure

without totally prohibiting harvest or consumption of the resource.

INTRODUCTION

The major components of a human health-risk assessment

are relatively well known. The “Red Book” of 1983 examined

how to assess risk to humans from cancer-causing

chemicals (NRC 1983). The paradigm set forth in this paper

has been expanded and modified over time, but the basic

principles have remained. To assess health risk from

chemicals, one must determine two things: the potential

effects of the chemical and exposure to it.

In order to estimate a chemical’s effect, the hazard and

dose/response of the substance must be examined. Hazard

is the inherent ability of a substance to cause harm. For

instance, a metabolite of benzo(a)pyrene (BaP), a

compound formed by the incomplete combustion of fossil

fuel, is thought to be a human carcinogen. So, the hazard of

the 7-8-dihydroxy-9-10-epoxide of BaP is that it can cause

cancer. Dose/response, on the other hand, estimates how

much of the substance is required, at the right place and

time, to cause a manifestation of the effect.

Once one knows what harm a chemical may inflict and how

much is required to do it, an estimate of exposure to that

chemical allows a characterization of the risk involved.

Action levels or tolerance levels are often defined as the

maximum concentration of a hazardous substance one can

have in food without experiencing the ill effects of the

substance. In effect, these are regulatory or riskmanagement tools used by the U.S. Food and Drug

Administration (FDA) or the U.S. Environmental Protection

Agency (EPA) to limit exposure to hazardous substances

from ingestion of contaminated food.

There are numerous difficulties involved in estimating the

hazard and dose/response of a chemical to humans. Are the

animal models used in toxicity tests appropriate for

extrapolating to humans? Is it a zero threshold chemical?

Are there sensitive subpopulations, etc.? There are also

numerous difficulties in estimating exposure to the

chemicals; these are often due to the inhomogeneous

nature of the substance in question in the environment.

Because a determination of risk can be no more accurate

than the most inaccurate number that goes into calculating it

(the Significant Figures Paradigm), one must strive to obtain

the very best estimates of all components of the assessment

process, including exposure.

Natural variability, a particularly difficult aspect of accurately

determining exposure to environmental contaminants, is the

focus of this manuscript. The much-studied chemical

contamination by kepone of the James River system is used

to exemplify various aspects of this phenomenon.

BACKGROUND

The James River is a major tributary to the Chesapeake Bay.

With its drainage basin of 25 600 km2, it delivers 16 percent

of the freshwater to the bay. The river is tidal for 160 km,

from its mouth near Norfolk, VA, to the fall line at Richmond,

VA. The salinity at its mouth is usually near 25 parts per

trillion (ppt), and brackish waters extend about 50 km

upstream, depending on precipitation within the drainage

basin. From this point upstream to Richmond, it is a tidal

freshwater river, and from this point to the mouth, it is

technically an estuary.

Approximately 110 km upstream, in the tidal freshwater

portion of the river, lies Hopewell, VA. From 1966 until 1974,
113

Figure 1—Histogram of kepone concentrations in bluefish from the

James River showing bimodal distribution.

the pesticide kepone was manufactured there by Allied

Chemical Company and Life Science Products, Inc. Over

this period, thousands of kg of kepone entered the river via

municipal waste effluents and other point and nonpoint

sources. Investigation of the river’s bottom sediments led to

an estimate of 14 000 kg of kepone being sorbed to the

particles. This became the major source of contamination to

the biota of the James River and the Chesapeake Bay

(Bender and Huggett 1984).

Kepone, (decachlorooctahydro-1, 3, 4-metheno-2Hcyclobuta (cd) pentelen-2-one), once intended to control fire

ants and cockroaches and eventually used for the banana

root borer in Central and South America, is extremely

persistent and bioaccumulative (Huggett and Bender 1980).

It is also a suspected human carcinogen.

Because kepone was found to be present in the edible

portion of finfish, crabs, clams, and oysters inhabiting the

James River and adjacent Chesapeake Bay, the EPA, under

the auspices of the Federal Insecticide, Fungicide,

Rodenticide Act, established action levels for the compound.

Based on NOELS from rodent testing and average food

consumption patterns, the following levels were originally

established: finfish: 0.1 mg per kilogram wet weight; crabs:

0.4 mg per kilogram wet weight; oysters: 0.3 mg per kilogram

wet weight. The action level for finfish was later raised to its

present level of 0.3 mg per kilogram.

It should be noted that action levels or tolerance levels are

enforceable by the FDA only if the contaminated commodity

crosses State lines. If the contaminated seafood, in this

case, was harvested and sold in Virginia, the Federal

government had no jurisdiction. Even so, the Commonwealth

of Virginia adopted the Federal action levels for instate

public health protection.

The problem that faced the State health regulators was to

determine accurately the concentrations that existed in the

seafood and to permit or ban fishing in certain areas or for

certain species in order to control exposure to kepone. This

required an accurate assessment of the kepone

concentration distributions in various species by location

and time. The findings of numerous sampling expeditions

and what is commonly called natural variability are

described and exemplified in the remainder of this paper.

KEPONE NATURAL VARIABILITY

Many, if not most, of the commercial species of finfish that

inhabit the Chesapeake Bay spend part of each year in the

ocean. In the spring, they re-enter the bay and spend

several weeks near its mouth, and hence, the James River,

to equilibrate to the lower salinities. Some organisms enter

the James River and remain throughout the season, and

others enter and leave again, migrating to other portions of

the bay. Concentrations of kepone in 91 blue fish collected in

June 1976 from one location in the bay are presented in

figure 1 (Huggett and others 1980). There is an obvious

biomodal distribution of concentrations with approximately

40 percent of the animals being above the action level of 0.1

mg per kilogram. These more contaminated animals

presumably spent some time in the James River before

leaving and grouping with fish that had not. Knowledge of

the distribution of concentrations greatly increases the

accuracy of estimates of exposure from consumption of the

fish.

Concentrations of kepone in the edible flesh of aquatic

organisms can also be influenced by the sex of the animal,

with males often being more contaminated. This is

presumably due to the substance being partitioned to the

lipid-rich eggs of the females (Huggett 1981). Figure 2

exemplifies this phenomenon. For the American shad, Alosa

sapidissima, 41 percent of the males contained more than

0.1 mg per kilogram whereas only 6 percent of the females

had levels as high.

Blue crabs, Callenectes sepidus, from the James River were

extensively sampled and analyzed. The difference in

concentrations between males and females was striking (fig.

2). As was the case with the American shad, the females

continued to show low levels relative to the males. All of the

male crabs sampled exceeded the action level of 0.4 mg per

kilogram whereas only 11 percent of the females were high.

This finding resulted in Virginia allowing commercial harvest

and sale of female crabs but not males.

The body burden of kepone is somewhat proportional to the

duration that the animal is exposed to the substance.

Croakers, Micropogon undulates, enter the bay from the

ocean in early spring and return in the fall. Croakers

collected from the James River throughout the summer of

1976 show that kepone concentrations increase with

exposure time with no plateauing noted (fig.3) (Huggett and

others 1980).

The physiology of the organism, as influenced by

temperature, can also affect tissue concentrations. This is

exemplified in figure 4, which shows kepone concentrations

in eastern oysters, Crassostrea virginica, from 1976 through

1980 (Bender and others 1985). During winter, when water

temperatures drop, the organisms have less planktonic food

available. Their water-filtering rate is drastically reduced, and

they metabolize stored body fat. In doing so, some of the

lipidophyllic kepone is liberated. As water temperatures rise,

feeding increases, and, concurrently, kepone levels

increase, thus giving rise to a yearly cycle of concentrations. Frequency 22

18

14

10

6

2

0

Kepone ppm

0.1 0.2 0.4 0.6 0.8 1.0

Fate and transport

Figure 1—Histogram of kepone concentrations in bluefish from the

James River showing bimodal distribution.
114

59% ≤ 0.1 ppm

94% ≤ 0.1 ppm

100% > 0.4 ppm

89% ≤ 0.1 ppm

Figure 3—Kepone concentrations in croakers, Micropogon

undulatus, collected from the James River in 1976.

Figure 2—Concentrations of kepone in male and female blue crabs, Callinectes sapidus, and American

shad, Alosa sapidissima.

1976 Action level = 0.1 ppm

6% > 0.1 ppm

 41% > 0.1 ppm

Male American Shad Female American Shad

1976 Action level = 0.4 ppm

11% > 0.4 ppm

Male Blue Crab Female Blue Crab

Croakers

Kepone (ppm)

Figure 4—Kepone residues in oysters vs. time.

 1976 1977 1978 1979 1980

James River

Oyster Rocks (moving average)

Kepone 

μg/g
115

Figure 4 also shows an overall decline in kepone

concentrations over time due to the burial of kepone in the

sediments after production of the pesticide was stopped in

1975.

Since the production of kepone stopped in the mid-1970s,

the concentrations in seafood have slowly decreased. The

harvest of finfish in the James River was totally or partially

banned until 1988 when concentrations fell below action

levels (Huggett 1989). Table 1 presents data that show that

by 1996 the levels were less than one quarter of the action

level (Unger, M.A. 2000. Personal communication. Glouster

Point, VA: Virginia Institute of Marine Science).

Table 1—Yearly averages of kepone concentrations in

striped bass (Morone saxatilis) from the James River

Animals Mean Standard

Year analyzed kepone deviation

- - No. - - Mg/kg

1989 40 0.13 0.03

1990 40 .20 .06

1991 40 .22 .07

1992 40 .20 .07

1993 40 .12 .11

1994 60 .17 .08

1995 60 .16 .07

1996 55 .07 .05

CONCLUSION

Studies designed to determine human exposure to

hazardous chemicals from consumption of contaminated

foods need to take into account natural variability. Failure to

do so may result in inadvertently placing consumers at risk

or, on the other hand, imposing harvest restrictions more

severe than necessary to protect public health. It is also

important to consider natural variability when assessing the

effects of hazardous chemicals on the organisms

themselves. Partitioning of lipidophyllic compounds to

gametes poses a relatively high risk of toxic reproductive

effects.

REFERENCES

Bender, M.E.; Huggett, R.J. 1984. Fate and effects of kepone in the

James River estuary. In: E. Hodgson, ed. Reviews in environmental toxicology. New York: Elsevier Science Publishers:

373–393.

Bender, M.E.; Huggett, R.J.; Slone, H.D. 1985. Kepone uptake: a

comparison of field and laboratory data. In: White, H.H., ed.

Concepts in Maine pollution measurements. College Park, MD:

Maryland Sea Grant Publications: 279–290.

Huggett, R.J. 1981. The importance of natural variability in the total

analytical scheme. Biomedical Mass Spectrometry. 8(9):

416–418.

Huggett, R.J. 1989. Kepone and the James River. In: Contaminated

marine sediments—assessment and remediation. Washington,

DC: National Research Council. National Academy Press:

417–424.

Huggett, R.J.; Bender, M.E. 1980. Kepone in the James River.

Environmental Science and Technology. 14(8): 918–923.

Huggett, R.J.; Nichols, M.M.; Bender, M.E. 1980. Kepone contamination of the James River estuary. In: Baker, R.A., ed. Contaminants and sediments. Ann Arbor, MI: Ann Arbor Science

Publishers. 1: 33–52.

National Research Council. 1983. Risk assessment in the Federal

government: managing the process. Washington, DC: National

Research Council. National Academy Press. 191 p.
116

Rising awareness of the important functions associated with

wetlands as well as the extent of historic losses have

stimulated societal perceptions of wetlands as rapidly

diminishing, yet, highly useful natural resources.

Consequently, public support (in principle) for restoration of

wetlands is quite high. However, a complex array of

socioeconomic factors as well as ecological uncertainties

continues to cause confusion in regard to implementation of

restoration efforts.

The term ecosystem restoration may vary in connotation

depending on the legal, political, scientific, or aesthetic

context in which it is used. We suggest that, regardless of

context, wetland restoration implies re-establishment of key

traits or functions that are valued by individuals and society

or both. Thus, the most common way to evaluate particular

restoration efforts is to compare functionality between

restored systems and a reference, which hopefully

represents the predisturbance state.

On former floodplains, afforestation of land previously used

for agriculture might be considered as wetland restoration.

Regeneration of tree seedlings on areas that were once

intensively cultivated may produce several effects that are

beneficial to society. These include reduction of sediment

export, improvement of wildlife habitat, production of wood

and fiber, and enhanced sequestration of carbon. However,

despite the value placed on these benefits by society, they

are not unique to wetlands.

It is arguable that the most important biogeochemical

function that some forest wetlands perform is filtration or

removal from waters of sediment, inorganic forms of

nutrients, and other substances considered impurities by

society. Obviously, this function is predominantly associated

with riverine systems, which develop and operate under a

lateral or flow-through type of hydrology so that water

continually passes through the wetland filter.

The biogeochemical filtration function of riverine forests is

closely linked with the inherently open nature of their

geochemical cycles. Nonriverine forest ecosystems as well

as those in which contact with their aquatic components has

been restricted tend to develop more closed geochemical

cycles as they aggrade. The degree of closure is associated

with the magnitude of outputs in relation to inputs. Whereas

an upland forest may theoretically aggrade to a point where

outputs or losses are negligible, i.e., a tight or closed cycle,

unaltered riverine forests are somewhat unique in that their

geochemical cycles do not approach this degree of closure.

The open-ended geochemical cycles of the riverine forests

allow exportation of significant quantities of dissolved

organic carbon (DOC), a critical source of energy for aquatic

food webs. This biogeochemical function is probably

secondary to filtration in terms of societal value but,

nonetheless, is quite important.

Consequently, many would agree that riverine forests, apart

from those physically separated from their river systems,

display two critically important biogeochemical functions:

filtration and DOC export. It is also apparent that both

functions are intrinsically linked with the flow-through

hydrology (and open geochemical cycles) that typifies

riverine systems. We suggest that these very simple ideas

can be used to examine the degree to which wetlands such

as bottomland hardwoods are being restored by particular

types of activities.

Afforestation of agricultural fields located on sites historically

occupied by riverine forests is currently a widespread activity

at the national as well as regional level (Clewel and Lea

POTENTIAL EFFECTS OF RESTORATION ON BIOGEOCHEMICAL

FUNCTIONS OF BOTTOMLAND HARDWOOD ECOSYSTEMS

Graeme Lockaby and John Stanturf 1

1 Lockaby, Professor, School of Forestry and Wildlife Sciences, Auburn University, Auburn, AL 36849; Stanturf, Soil Scientist, Center for

Bottomland Hardwoods Research, Southern Research Station, P.O. Box 227, Stoneville, MS 38776, Present address: Forestry Sciences

Laboratory, 320 Green St., Athens, GA 30602-2044.

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability of

wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.

Abstract—The concept of wetland restoration carries multiple meanings and implications. The scientific usage of the term

connotes re-establishment of wetland functions, and often it is the functions, which society deems most valuable, that

receive highest focus. Arguably, among key wetland functions, the highest societal value may be linked with the biogeochemical transformation or filtration function, a key contributor to maintenance of water quality. This function requires

flow-through hydrology such as that associated with unimpeded riverine forests, and, consequently, its re-establishment is

negated by the absence of such hydrology. Consequently, afforestation of former agricultural areas, which were protected

from flow-through hydrology, i.e., flooding, by dikes, ditches, etc., cannot be considered restoration in a complete sense

unless some semblance of flow-through hydrology is also restored. The term quasi-depressional wetland is suggested as

being appropriate for afforested areas where hydrologic restoration is unfeasible.
117

1990; Sharitz 1992; Stanturf and others 1998, 2000). In the

Mississippi Delta region of the Southern United States, such

afforestation is often referred to as bottomland hardwood

restoration (King and Keeland 1999, Stanturf and others

2000b). As previously mentioned, there are several excellent

reasons for pursuing such activities. However, the topic at

hand dictates that we examine this practice from a

biogeochemical standpoint.

In the vast majority of these cases, the practice of

agriculture on floodplains necessitated protection from

annual flooding in addition to removal of forests. Most often,

dikes, levees, or ditches or all were used to physically

separate the aquatic and terrestrial components. Obviously,

these structures were intended to modify or eliminate the

flow-through hydrology that typifies riverine forests. In most

cases following agricultural abandonment, socioeconomic

factors cause removal of the protective structures to be

unfeasible.

An exception is the major restoration effort associated with

the Kissimmee River and adjoining lands in Florida (http://

www.saj.usace.army.mil/ dp/kissimmee.html). In that case,

approximately 100,000 acres of floodplain are being

restored to a state subject to natural flooding. However, as

would be expected, restorations at this scale and degree of

complexity are very costly, i.e., $372 million (http://

www.fcn.state.fl.us/eog/govdocs/opbenv/ saveglades/

everglades/html/kissimee.htm).

It should be noted that some afforestation areas retain flow

through hydrology. These include areas within the batture

(unprotected floodplain of the Mississippi River) or those

tributaries such as the Yazoo subject to backwater flooding.

However, as previously mentioned, restoration of natural,

riverine flooding regimes is rarely feasible. As a result, the

aggrading forests on afforested sites cannot exhibit openended biogeochemistry and, consequently, will not function

as biogeochemical filters or exporters of DOC. This limitation

of afforestation activities has been previously recognized

(Allen 1997, King and Keeland 1999). Suggested remedies

have included plugging drainage ditches or building water

control structures on portions of the afforested sites so that

controlled flooding can be induced in much the same way

that it is applied within greentree reservoirs. On public land

such as national wildlife refuges and national forests,

relatively large areas have been restored in this fashion as

greentree reservoirs, moist soil management units, or

permanent water bodies. In addition, it is not uncommon for

some flooding to occur on lower lying portions from

accumulation of precipitation.

However, it should be recognized that these types of

flooding reflect those more commonly associated with

depressional wetlands than with riverine. Since it is

commonly acknowledged that hydrology represents the

dominant controlling process within a wetland (Mitsch and

Gosselink 1993), the nature of the hydrology in the restored

system is a critical factor in our evaluation of restoration

success. Thus, the significance of restoring a former riverine

system to quasi-depressional hydrology lies in

biogeochemical differences between depressional vs.

riverine wetlands.

The chief contrast between riverine vs. depressional

hydrology is a predominance of lateral vs. vertical flows

respectively (Brown 1990). Consequently, biogeochemical

functions are aligned primarily with precipitation,

evapotranspiration, and infiltration in basins (fig. 1) as

opposed to sheet flow in riverine systems (fig. 2). Lugo and

Figure 1—Hydrologic budget for a typical depressional wetland, a cypress dome

in Florida. Units are in centimeters per year; depressional wetlands typically

have greater inflow than outflow of surface water (Brown 1990).

51 0

58

107 127

S = - 28
118

others (1990) felt that this hydrologic distinction drives

differences in terms of how elements are accumulated in the

two wetland types as well as which elements would be

limiting to NPP; e.g., theoretically P would be more limiting in

basins due to the scarcity of vertical input mechanisms.

However, we should also be cautious in assuming that this

type of afforestation will result in the formation of typical

basin wetlands. There likely will be functional distinctions

between a basin wetland, which formed and functions under

true basin hydrology, vs. an afforested field, which may

occasionally accumulate water. We suggest use of the term

quasi-depressional system to indicate this distinction.

Because these quasi-depressional systems remain isolated

from riverine influences, they contribute little to

biogeochemical filtering or to DOC export to aquatic

systems.

It is apparent that some basin wetland forests are capable of

accumulating nutrients when subjected to elevated inputs

(Ewel and Odum 1984). There is also evidence that

infiltration export is generally low (Brown 1990), and,

consequently, some basins are regarded as nutrient

accumulation zones rather than filters or exporters of DOC.

However, their lack of hydrologic linkages to major

landscapes minimizes their biogeochemical value to society.

In particular, the quasi-depressional systems created behind

levees or dikes may have even less opportunity to

accumulate nutrient inputs because they may exist in a more

hydrologically isolated state than true basins.

The U.S. Environmental Protection Agency (EPA) has

identified the Yazoo-Mississippi Basin as an area of

significant concern for surface and ground water quality.

Although surface water runoff in the Lower Mississippi

Alluvial Valley (LMAV) contributes only 20 percent of the

nitrate loading implicated in the expansion of the hypoxic

zone in the Gulf of Mexico, the EPA is expected to focus

significant resources on the LMAV to improve water quality.

Policy alternatives under consideration include reducing

nitrogen use by 20 to 40 percent and converting agricultural

land to forests in an effort to restore and enhance natural

denitrification processes (EPA 1999). The assumption is

made that restoration (afforestation) of bottomland hardwood

forests will reduce nutrient export into the Gulf. This will be

true to the extent that a potential source of nutrients will be

reduced by changing land use from row crop agriculture to

forests (Thornton and others 1998). But the restored system

will play at most a small role as a nutrient filter unless it is

hydrologically linked to a riverine system. Thus, the greater

benefit, in terms of nutrient filtration, would come from

conversions within the active floodplain of small rivers

throughout the basin of nutrient origin and from buffer strips

planted along drainage ways (Castelle and Johnson 2000,

Castelle and others 1994). Whereas forested buffer strips

may provide advantages to the landowner over grass or

herbaceous strips, the relative effectiveness of forest vs.

grass buffers on nutrient filtration remains uncertain.

Whereas public support for wetland restoration has

traditionally been strong, that support may weaken when

local populations are made aware of the implications of true

hydrologic restoration within a former floodplain. Such

restoration will almost certainly necessitate elimination of

land uses that are incompatible with significant flooding,

e.g., farming, habitation, etc. Thus, added to the engineering

cost of levee removal, etc., are significant compensation

expenditures to landowners. In addition, the unpopular

105 72

229

22 21

232 S = 0 5,300

Flooding

River

Figure 2—Hydrologic budget for a typical riverine wetland, an alluvial cypress swamp in Illinois.

Units are in centimeters per year; riverine wetlands typically exhibit flow-through hydrology (Mitsch

and Gosselink 1993).
119

nature of a major taking of private property will also weaken

political support for State and Federal appropriations to

support the projects. While we neither advocate nor support

this approach, the obstacles associated with such an

endeavor appear to be significant, and, thus, will make

implementation of true riverine restorations difficult to

achieve.

In summary, it is recognized that there are many very

worthwhile reasons for conversion of former farmland to

hardwood forests. However, continued growth in population

levels dictates that utilization of landscapes will grow

increasingly complex in the years to come. Similarly, it will

become critically important that professionals as well as the

general public be aware of the specific nature of the

changes that are induced (or not induced) by farmland

conversion. Only the development and widespread

dissemination of that knowledge will protect society from

unrealistic expectations, which lead to landscape-level

mistakes in natural resource management.

REFERENCES

Allen, J.A. 1997. Reforestation of bottomland hardwoods and the

issue of woody species diversity. Restoration Ecology. 5(2): 125–

134.

Brown, S. 1990. Structure and dynamics of basin forested wetlands

in North America. In: Lugo, A.E.; Brinson, M.; Brown, S., eds.

Ecosystems of the world 15: forested wetlands. New York:

Elsevier: 171–199.

Castelle, A.J.; Johnson, A.W. 2000. Riparian vegetation effectiveness. NCASI Tech. Bull. 799. Research Triangle Park, NC:

National Council for Air and Stream Improvement, Inc. 26 p.

Castelle, A.J.; Johnson, A.W.; Conolly, C. 1994. Wetland and stream

buffer size requirements—a review. Journal of Environmental

Quality. 23(5): 878–882.

Clewell, A.F.; Lea, R. 1990. Creation and restoration of forested

wetland vegetation in the Southeastern U.S. In: Kusler, J.A.;

Kentula, M.E., eds. Wetland creation and restoration. Wetland

creation and restoration: the status of the science. Washington,

DC: Island Press: 195–231.

Dahm, C.N.; Cummins, K.W.; Valett, H.M.; Coleman, R.L. 1995. An

ecosystem view of the restoration of the Kissimmee River.

Restoration Ecology. 3: 225–238.

Ewel, K.C.; Odum, H.T. 1984. Cypress swamps. Gainesville, FL:

University Presses of Florida. 472 p.

King, S.L.; Keeland, B.D. 1999. Evaluation of reforestation in the

Lower Mississippi River Alluvial Valley. Restoration Ecology. 7:

348–359.

Lugo, A.E.; Brinson, M.M.; Brown, S. 1990. Synthesis and search for

paradigms in wetland ecology. In: Lugo, A.E.; Brinson, M.;

Brown, S., eds. Ecosystems of the world 15: forested wetlands.

New York: Elsevier: 447–460.

Mitsch, W.J.; Gosselink, J.G. 1993. Wetlands. 2d ed. New York: Van

Nostrand Reinhold. 722 p.

Sharitz, R.R. 1992. Bottomland hardwood wetland restoration in the

Mississippi drainage. In: Restoration of aquatic systems:

science, technology, and public policy. Washington, DC: National

Academy Press: 496–605.

Stanturf, J.A.; Schoenholtz, S.; Schweitzer, C.J.; Shepard, J.P. 2001.

Achieving restoration success: myths in bottomland hardwood

forests. Restoration Ecology. 9(2): 189–200.

Thornton, F.C.; Joslin, J.D.; Bock, B.R. [and others]. 1998. Environmental effects of growing woody crops on agricultural land: firstyear effects on erosion and water quality. Biomass and

Bioenergy. 15: 57–69.

Toth, L.A. 1995. Principles and guidelines for restoration of river/

floodplain ecosystems—Kissimmee River, Florida. In: Cairns, J.,

ed. Rehabilitating damaged ecosystems. 2d ed. Boca Raton, FL:

Lewis Publishers: 49–73.

Toth, L.A.; Anderson, D.H. 1998. Developing expectations for

ecosystem restoration. In: Wadsworth, K.G., ed. Transactions of

the 63rd North American wildlife and natural resources conference; 1998 March 20–24; Orlando, FL. Washington, DC: Wildlife

Management Institute: 122–134.

U.S. Environmental Protection Agency. 1999. Integrated assessment

of hypoxia in the northern Gulf of Mexico. Stennis Space Center,

MS: U.S. Environmental Protection Agency, Gulf of Mexico

Program, Stennis Space Flight Center. 54 p. [Draft for public

comment].
120

INTRODUCTION

This paper summarizes the context of forest restoration in

the Nordic countries and Estonia including the debate on

changing aims of future forestry and silviculture.

Furthermore, we present the objectives and the main

conclusions of a cooperative project between the Nordic

countries, Estonia, and the United States on direct seeding

of broadleaves. The cooperative project was supported from

both national funds from each of the participating countries

as well as funds from the Nordic Forest Research

Cooperation Committee (SNS).

CONTEXT OF FOREST RESTORATION

The Nordic countries include Iceland, Norway, Sweden,

Finland, and Denmark, which range in latitude from 54° N in

southern Denmark to 72° N at North Cape, Norway. Estonia

is one of the Baltic countries and is located south of Finland

(fig. 1). Forest cover ranges from a very limited part of the

land area in Iceland to a dominant part in Sweden and

Finland (table 1).

The boreal coniferous vegetational zone dominates the area.

Denmark and southern Sweden are, however, located in the

deciduous (nemoral) forest zone, whereas the interior part of

Iceland and the high altitudes of Norway and Sweden are in

the mountainous zone (Walter 1985). Forest industry plays

an important role for the Swedish, Finnish, Norwegian, and

Estonian economy, but the economical importance of wood

production is rather marginal for Denmark and Iceland. In

general, most of the forestland is private property, which is a

relatively new ownership status in Estonia.

The term forest restoration covers very different silvicultural

challenges in the Nordic countries. In Iceland, afforestation

on totally barren and degraded land, practically deserts, is

one important approach to forest restoration. Special

attention is paid to restoration of the birch woodlands, which

covered more than 25 percent of the land area at the time of

settlement in the 10th century (Sigurdsson 1977). In contrast

to the Icelandic situation, afforestation efforts in the other

Nordic countries and Estonia occur on fertile farmland. Aims

of afforestation are rather different within and between the

countries. In Finland, Sweden, and Norway the expected

extent of afforestation is rather limited (table 1) and serves

mainly as an alternative land use to small scale, inefficient

agriculture. In Estonia, many small farms have been turned

over to the descendants of former owners. They have no

experience and knowledge of farming practices, so forestry

may be of interest to these landowners as a low-cost, landuse alternative. Consequently, a significant increase in

forestland on abandoned farmland is expected in Estonia. In

Denmark, the goal of the afforestation program is to double

the forested area within one tree rotation (about 100 years).

There are several aims of this program including:

• increased concern for sustainability, nature conservation,

and biodiversity;

• protection of ground water resources;

FOREST RESTORATION IN THE NORDIC COUNTRIES

Palle Madsen, Ása L. Arad•ttir, Emile Gardiner, Pelle Gemmel, Kåre Lund Høie,

Magnus Löf, John Stanturf, Peter Tigerstedt, Hardi Tullus, Sauli Valkonen, and Veiko Uri1

1 Madsen, Senior Researcher, Danish Centre for Forest, Landscape, and Planning (FSL), Kvak Møllevej 31, 7100 Vejle, Denmark; Arad•ttir,

Division Leader, Soil Conservation Service, Gunnarsholt, 850, Hella, Iceland; Gardiner, Research Forester, and Stanturf, Project Leader, Center

for Bottomland Hardwoods Research, Southern Research Station, P.O. Box 227, Stoneville, MS 38766, Present address: Forestry Sciences

Laboratory, 320 Green St., Athens, GA 30602-2044; Gemmel, Professor, and Löf, Researcher, Swedish University of Agricultural Sciences, S230-53, Alnarp, Sweden; Høie, Senior Researcher, The Norwegian Crop Research Institute, Plant Protection Centre, N-1432 Ås, Norway;

Tigerstedt, Professor, University of Helsinki, FIN-00014 Helsinki, Finland; Tullus, Professor, and Uri, Researcher, Estonian Agricultural

University, Tartu, Estonia; Valkonen, Senior Researcher, Finnish Forest Research Institute, FIN-01301 Vantaa, Finland.

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability of

wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.

Abstract—The Nordic countries include Iceland, Norway, Sweden, Finland, and Denmark, which range from lat. 54° in

southern Denmark to lat. 72° at North Cape, Norway. This region is dominated by the boreal coniferous vegetational zone.

Denmark and southern Sweden are, however, located in the deciduous (nemoral) forest zone, whereas the interior part of

Iceland and the high altitudes of Norway and Sweden are in the mountainous zone. Forests cover 1, 11, 28, 60, and 66

percent of the land area in Iceland, Denmark, Norway, Sweden, and Finland, respectively. Traditional forestry has mainly

concentrated on conifers in the boreal and the nemoral zones. Increased concern about nature conservation and sustainable land use along with economic constraints and reduced softwood timber prices have led to increased focus on growing

and regenerating broadleaves. Planting nursery-grown seedlings is the common and preferred method for forest restoration

and afforestation. However, this technique is expensive unless wide spacing is used. Direct seeding is a less expensive

alternative that may still give acceptable results. This regeneration practice has been examined for a number of broadleaved

species within a collaborative project including forest researchers from Mississippi, Estonia, and the five Nordic countries.

Results from this work are presented, and conclusions are drawn for forest management and future research and development in this field.
121

• improvement of recreational values of the landscape;

and

• reduction of subsidized agricultural production.

Moreover, at present in the deciduous zone (southern

Sweden and Denmark), conifer plantations are being

transformed into broadleaf stands, particularly on better

soils. Today, conifer plantations cover about 65 to 80 percent

of the forestland in this area, and the main species is

Norway spruce [Picea abies (L.) Karst]. Norway spruce is

outside or on the edge of its natural range in Denmark and

southern Sweden. The conifers were initially planted due to

their high productivity and low cultivation costs. However,

they have shown in many cases poor wind stability and

health depending on site and species. Such catastrophes

not only destroy the existing forests but also leave an area

with considerable regeneration problems. The forest climate

is lost, and weeds, frost, drought, and wind may cause

problems for regeneration.

As illustrated above, the background for forest restoration

with respect to ecological conditions and forest industry is

very diverse in the Nordic countries. However, cultural,

political, and economical similarities of the countries form a

common platform for changes and aims in the forestry of

these countries.

Table 1—Forestland and woodlands in the Nordic countries and Estoniaa

Forestland and Forestland, % Expected afforestation, Private- (includ. companies)

Country woodlands of total land area % of total land area owned forestland

- - 1,000 km2 - - - - - - - - - Percent - - - - - - -

Finland 201 66 1 71

Sweden 244 60 1 78

Estonia 21 50 7 35

Norway 87 28 < 1 85

Denmark 5 11 11 69

Iceland 1 1 2–5 70

a Woodlands are roughly nonproductive in terms of wood production. In Iceland, in particular, they cover a significant part of the forested land.

Expected afforestation is estimated based on personal judgements for the next 50 to 100 years (Finland, Sweden, Norway, Iceland, and

Estonia), or it is based on a political decision (Denmark). In Estonia, the government aims at a 50-percent private forestland rate.



 





 







   



   



    







     



 









  

 







 



 





 









































  



 



























 





 



 





 

  







 



























 





 











 

 





















 



 





 

 





 

    

 





 



 





 

 





 



   









 

 







 















Sweden

Denmark

Estonia

Finland

Norway

Iceland

Figure 1—The Nordic countries and Estonia.
122

CHALLENGES IN SILVICULTURE

Traditional forestry in the Nordic countries has concentrated

mainly on growing conifers for timber and pulp in the boreal

and the nemoral zones. During the past two decades,

increased concern over ecological sustainability, nature

conservation, and sustainable land use in conjunction with

economical constraints and reduced softwood timber prices

has led to an increased focus on the use of broadleaf

species and close-to-nature forest management. The more

diverse and multifunctional aims of forestry have

emphasized the need for the forests to be flexible with

respect to future outputs as wood and nonwood products

and values. Additionally, the importance of such flexibility is

stressed by the long production periods in Nordic forestry.

The rotation length usually ranges between 50 to 120 years,

depending on site, species, silvicultural practices, etc. The

problem is that the main role of the future forests cannot be

predicted precisely.

The stability of forests can be expressed in terms of

resistance and resilience of the forest ecosystem (Larsen

1995). Poor resistance may express a considerable

susceptibility of the forest to be damaged or destroyed by

strong winds, drought, fire, or a complex of factors. Poor

resilience may entail considerable regeneration problems

because of the difficulties regaining the forested condition

after a catastrophe. Active forest management is crucial for

maintaining a stable forest ecosystem. Otherwise, there is

considerable risk that the forest will not fulfil the aims of its

establishment.

Traditional and close-to-nature forest management is

probably best described as relative degrees of forest

management, which may overlap. Stands resulting from

traditional forest management are usually homogeneous

with respect to species and age. In traditional forest

management, stands are harvested when the average tree

has reached maturity, and planting is the dominant

regeneration method with natural regeneration utilized to a

lesser extent. Simple administration, apparently costeffective methods, and convincing economical models are

probably the main reasons for the widespread use of this

forest-management approach.

In close-to-nature forest management, natural forest

ecosystem processes are used and supported by the

silviculturist to achieve inexpensive natural regeneration and

optimise production value of each tree during the later part

of the rotation period. The latter is performed by harvesting

single trees as they reach a target diameter; e.g., 57 cm

d.b.h. for beech (Fagus sylvatica L.) in Denmark and

southern Sweden. Close-to-nature forestry is also

characterized by the high priority given to the use of siteadapted species in heterogeneous forest structures, which

will gradually develop as a consequence of the regeneration

system. It should be stressed that stand heterogeneity is not

the goal itself but a way to allocate the species to various

soil conditions and improve forest-floor conditions for natural

regeneration.

Proponents of the principles of close-to-nature forest

management regard it as a silvicultural system that may

support the multifunctionality of a forest. This may include

economic benefits from the extensive use of natural

regeneration and single-tree management.

WHY IS DIRECT SEEDING INTERESTING?

In close-to-nature forestry, how is the strong focus on

minimizing cultivation costs and promoting site-adapted

species relevant to forest restoration? Low cultivation costs

may be essential for the long-term economic success of

close-to-nature silviculture. Natural regeneration is, in many

cases, not a reliable and realistic option on sites that lack

relevant seed sources or appropriate germination

environments. Wind-dispersed species like birch (Betula spp.

L.) may, however, have a potential for natural regeneration

on bare land. Examples of successful natural regeneration of

birch on former farmland or on clear-felled areas are well

known. It was common practice in traditional forestry to

remove regenerated birch to reduce competition on planted

conifers. In Iceland, it is proposed to plant birch at very wide

spacings or in clusters to distribute seed sources in the

landscape and thereby support the restoration of the birch

woodlands (Arad•ttir 1991).

Planting seedlings is the most common practice of today’s

forest restoration programs. It is a reliable but expensive

method, particularly with respect to establishment of

broadleaves. It typically costs $0.40 to $0.70 to purchase

and plant seedlings. Usually 5,000 broadleaf seedlings are

planted per hectare. Direct seeding is an alternative

regeneration method that may be less costly. For $1.00 it is

possible to buy 20 to 100 acorns or 100 to 200 beechnuts;

but, in bumper crop years, the prices may decrease.

Consequently, higher seedling densities could potentially be

established with direct seeding at lower costs than planting

seedlings. High-density stands could also have a positive

impact on future stand quality. Table 2 outlines the

approximate ranges of costs related to direct seeding and

planting of beech and oak. The oak may be either

pedunculate oak (Quercus robur L.) or sessile oak (Q.

petraea Liebl.). The question is whether it is possible to

Table 2—Approximate costs for direct seeding of beech

and oak in Denmark and southern Swedena

Direct seeding Planting

 - - - - Dollars per hectareb - - - -

Site preparation 0 – 700 0 – 700

Seeds or transplants 200 – 500 1,200 – 2,400

Sowing or planting 100 – 350 400 – 800

Fence 0 – 1,300 0 – 1,300

Total 300 – 2,850 1,600 – 5,200

a The costs depend to a large extent on several factors such as

management objectives, economy of the landowner, goals for future

wood quality, site quality, deer population, area of the site, and cost

of seed/transplants.

b United States currency.
123

achieve the knowledge, skills, and methods that make direct

seeding competitive with planting in terms of costs,

reliability, and stand quality (Johnson 1981, Kübner and

Wickel 1998, Leder and Wagner 1996). Additionally, direct

seeding may offer a number of other advantages compared

to planting:

• The natural root development of seedlings that developed

from direct seeding may be advantageous with respect to

future stability against windthrow or drought events.

• The high stock density may reduce the need for deer

fences.

• Direct seeding of species mixtures could be a relatively

inexpensive way to establish a stand with good potential

for structural adaptation to microsite variability.

Governmental subsidies to establish broadleaves are

commonly available for foresters and forest authorities in

Sweden and Denmark. Subsidies in Denmark (table 3) are

supposed to support the fulfilment of the Danish

afforestation program and enhance the transformation of

conifer plantations into hardwood forests. Subsidies are,

however, not necessarily prudent in terms of economical

sustainability of forestry, and they call for further research

and development that eventually can provide forestry with

more inexpensive regeneration methods.

CURRENT PRACTICE

There is limited current knowledge and experience on

direct seeding broadleaves in the Nordic countries.

However, direct seeding was commonly practiced to

establish beech and oak in southern Sweden and Denmark

during the late 19th and the early to mid-20th century. The

intensity of these former regeneration efforts far exceeded

the present level in terms of labor input, and today it seems

relevant to draw an analogy to horticulture. Intensive

weeding and pest-management practices were applied

together with the use of very high seed densities. Stock

densities exceeding 100,000 seedlings per hectare were

not unusual. Regulations on provenance use reduced-seed

availability and consequently led to the increased use of

nursery stock.

Today, direct seeding of oak is gaining new popularity for

afforestation of farmland in southern Sweden and Denmark.

The method is regarded as reliable, and costs are between

30 to 50 percent of costs for planting seedlings. Oak seeding

can be reliable on bare farmland because of the absence of

small rodents, as they have no access to vegetative cover.

Moreover, European oaks typically exhibit pioneer

characteristics, which make them well adapted to the site

conditions on open fields.

DIRECT SEEDING PROBLEMS

At present, there are a number of practical reasons for the

limited use of direct seeding. Insects, slugs, rodents, birds,

and deer can consume large amounts of seeds and young

seedlings (Nielsson and others 1996). Late-spring frost may

damage sprouting seedlings, and germination can fail or be

delayed due to drought or incompletely broken seed

dormancy. Moreover, seed availability is limited by protocols

established for seed-source approval. These limiting factors

must be addressed if direct seeding is going to be a planting

alternative.

Furthermore, it is difficult to successfully incorporate new

methods into silvicultural practice. Increasingly, many

foresters face time constraints in their work, which reduces

the time available for learning new regeneration methods.

DIRECT SEEDING RESEARCH AND

DEVELOPMENT

The ultimate goal of the joint Nordic project on direct

seeding of broadleaves was to develop new, reliable, and

inexpensive regeneration methods compared to the

conventional practice of planting seedlings. Furthermore, it

was the aim of the project to test new methods on a range of

typical site types for forest restoration.

The main hypothesis was that seedlings could be

established through direct seeding with similar success as

seen for planted nursery seedlings if:

• seed was of high quality;

• seed dormancy was broken by a relevant pretreatment

procedure before sowing; and

• seed and established seedlings were protected against

pests.

Table 3—Examples of present subsidy programs for

afforestation and regeneration in Denmarka b

Maximum

Forest restoration type subsidy

Afforestation in areas where

it is highly preferred (per ha)

Planting of broadleaves $2,900

Planting of conifers 1,900

Direct seeding 1,900

Afforestation in other areas (per ha)

Planting of broadleaves 1,900

Planting of conifers 1,200

Direct seeding 1,200

Additional subsidies for afforestation

Pesticide-free afforestation (per ha) 500

Fence (per m) 2

Income compensation (per ha per yr) 350

Regeneration on forest land (per ha)

Natural regeneration of beech, oak, or ash

(Fraxinus excelsior) 1,300

Planting beech, oak, ash, or basswood

(Tolia cordata or T. platyphyllos) 3,000

a Eight to twelve years after establishment there must be a

minimum of 2,500 to 4,000 saplings (depending on species and site

type) per hectare with an average height of more than 1 m.

b United States currency.
124

Additionally, it was hypothesised that short seeding tubes

(10 to 25 cm tall) could protect seed and seedlings

sufficiently. Light-seeded species like birch and alder (Alnus

spp. Ehrhart) were, however, not expected to benefit from

the protection against rodents. Instead, the tubes were

supposed to create a calm and moist microclimate. This

would hypothetically prevent light seeds from being blown

away, improve seed germination, and improve initial seedling

growth.

WHAT WAS INVESTIGATED?

Approximately 50 field experiments were established in the

participating countries from spring 1995 until spring 1998.

As many as three experiments were carried out at the same

site in successive years. The main emphasis in the Nordic

countries was generally on pedunculate oak, beech, and

birch (Betula pubescens Ehrh. or B. pendula Roth) (table 4),

but other species were also tested. In Mississippi, water oak

(Q. nigra L.) was investigated, and in Estonia, grey alder

(Alnus incana Moench), pedunculate oak, and birch were

investigated. In Denmark, a number of other species were

tested including wild cherry (Prunus avium L.), hawthorn

(Crataegus monogyna Jacq.), ash (Fraxinus excelsior L.)

and sycamore maple (Acer pseudoplatanus L.).

Various types of short seed tubes were tested. The tubes

were manufactured in Denmark especially for this project,

and their design was changed from year to year after

evaluation of preliminary results. Initially (1995), the tubes

were 25 cm tall, 28 mm in diameter, and were made out of

polyethylene. In 1996 tubes of biodegradable plastics

(mainly a mix of cellulose and starch) were used. Tube

lengths and diameters tested ranged from 5 to 25 cm and 14

to 38 mm, respectively.

There were generally two types of experiments:

• Some experiments were relatively intensive and included

both planted seedlings and sown seed; the latter seeded

with and without seed-tube protection. These experiments

were intensively monitored. Some of these intensive

studies also included treatments with different densities

of a lupine (Lupinus nootkatensis Donn ex Sims) cover

crop (Iceland), weed control treatments (Norway,

Denmark, Sweden, Estonia), or soil preparation

(Sweden).

• Other experiments were nonintensive and were often

established in close collaboration with forest managers

(Denmark, Finland). These low-budget trials served as a

supplement to the more intensive experiments allowing

direct seeding and the seed tubes to be tested at a range

of sites in different years.

CONCLUSIONS OF THE DIRECT SEEDING

PROJECT

Results are not presented here since some of the work is

still in the process of being published. However, some of the

main research conclusions, implications for forest

management, and directions for future research are

summarized.

SEED TUBES

Seed tubes showed promising preliminary results with both

light-seeded and heavy-seeded species. In many cases,

germination was good in the tubes, and they protected the

seed and seedlings from rodents and weevils. However, it

was not complete protection, and, in some cases, the tubes

even increased rodent problems because the animals

learned that the tubes contained food. Additionally, some

Table 4—Main species and site types of the experiments in the participating countries

Country Main species Site types

Iceland Birch (Betula pubescens) Severely disturbed or partially reclaimed soils,

which are being colonized by lupines

Norway Birch (B. pendula) Agricultural fields

Sweden Beech (Fagus sylvatica) and

pedunculate oak (Quercus robur) Clearcut following a conifer stand, conifer shelterwood, or

agricultural fields

Denmark Beech (F. sylvatica) and

pedunculate oak (Q. robur) Clearcuts following conifer stands, conifer or broadleaved

shelterwoods, or agricultural fields

Finland Pedunculate oak (Q. robur) Agricultural fields

Estonia Birch (B. pendula) Agricultural fields or a clearcut following a conifer stand

United States

Mississippi Water oak (Q. nigra) Agricultural fields
125

seedlings suffered from winter frost damage. The warm and

moist microclimate inside the seed tubes apparently

prevented some seedlings from developing frost hardiness.

Other problems encountered included waterlogging in the

tubes, poor light conditions caused by soil sticking to the

tubes, and frost heaving of the tubes. Moreover, some of the

biodegradable materials used for tube construction

apparently decreased germination of birch, alder, and

beech. In summary, the seed tubes tested in these

experiments did not lead to new, reliable, and inexpensive

regeneration methods.

FOREST MANAGEMENT

In spite of the shortcomings of the seed tubes, we did obtain

relevant knowledge that can be applied towards future forest

management, research, and development. On the degraded

sites of Iceland, planting gave higher rates of birch

establishment than did direct seeding, but the seeding

success varied between sites and appeared promising on

some sites. Furthermore, a properly managed lupine cover

crop improved birch establishment. On farmland in Norway

and Estonia, no reliable methods were found for direct

seeding birch and alder. In these countries, weed

competition was identified as a main problem for the small

seedlings.

The research indicated that heavy seeded species, like

beech and oak, may be successfully established by direct

seeding on sites and in years with low rodent populations.

However, competitive weeds may still need to be controlled,

particularly on fertile soils. Additionally, site suitability of

each species must be considered for inexpensive and

successful regeneration. For example, beech favors the

environmental conditions under a shelterwood, whereas oak

is better adapted to the microclimate of clearcuts or

farmland.

FUTURE RESEARCH AND DEVELOPMENT

The Nordic project on direct seeding broadleaves has

identified the need for further development of silvicultural

skills, methods, and direct-seeding techniques.

Additionally, there is a need for monitoring methods and

rodent population control in concurrence with seeding of

heavy seeded tree species. Likewise, the development of

pretreatment methods aimed particularly at sites to be

direct seeded is regarded as an important field of

research.

REFERENCES

Arad•ttir, A.L. 1991. Population biology and stand development of

birch (Betula pubescens Ehrh.) on disturbed sites in Iceland

[Abstract]. College Station, TX: Department of Range Science,

Texas A&M University. Ph.D. dissertation. [In: Dissertation

abstracts international. 52–06B: 2,869].

Johnson, R.L. 1981. Oak seeding—it can work. Southern Journal of

Applied Forestry. 5(1): 28–33.

Kübner, R.; Wickel, A. 1998. Entwicklung einer Buchensaat (Fagus

sylvatica L.) unter Fichte (Picea abies (L. Karst.)) im

Osterzgebirge. Forstarchiv. 69: 191–198.

Larsen, J.B. 1995. Ecological stability of forests and sustainable

silviculture. Forest Ecology and Management. 73: 85–96.

Leder, B.; Wagner, S. 1996. Bucheckern/Streu- Voraussaat als

Alternative beim Umbau von Nadelholzreinbeständen in

Mischbestände. Forstarchiv. 67: 7–13.

Nielsson, U.; Gemmel, P.; Löf, M.; Welander, T. 1996. Germination

and early growth of sown Quercus robur L. in relation to soil

preparation, sowing depths, and prevention against predation.

New Forests. 12: 69–86.

Sigurdsson, S. 1977. Birki á Íslandi (Birch in Iceland). In:

Gudmundsson and others, eds. Skógarmá. Reykjavik, Iceland:

Edda Print: 146–172. In Icelandic.

Walter, H. 1985. Vegetation of the earth and ecological systems of

the geo-biosphere. 3rd ed. New York: Springer-Verlag. 318 p.
126

APPLICATION OF THE SOIL PERTURBATION INDEX TO EVALUATE CREATED

AND RESTORED WETLANDS

Rebecca Smith Maul and Marjorie M. Holland1

1

 Maul, Graduate Student, Department of Geology and Geological Engineering, The University of Mississippi, University, MS; Holland,

Associate Professor, Department of Biology, The University of Mississippi, University, MS 38677.

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability

of wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.

Abstract—Biogeochemical properties of wetlands have recently been investigated to assess recovery of wetland ecosystems following human alteration. Analyses of soil samples have shown that the natural regeneration of timber-harvested

wetlands exhibits predictable trends for soil organic matter, total organic carbon, total Kjeldahl nitrogen, and total phosphorus. Incorporating these four nutrients, a Soil Perturbation Index (SPI) was previously developed to aid in biogeochemical

comparisons of altered wetlands at different successional stages to mature reference wetlands. The current study explores

whether reforested sites previously in agriculture exhibit similar biogeochemical trends to forested wetlands previously

timber harvested. Results indicate that reforested sites previously in agriculture exhibit biogeochemical trends similar to

timber-harvested wetlands, although perhaps at slower rates. Trends for the SPI developed from the agriculturally based

sites were very similar to that of the original SPI developed from naturally regenerated sites.

INTRODUCTION

The Nature Conservancy (1992) estimates that about 4.9

million ha of forested wetlands remain in the Lower

Mississippi River Alluvial Valley. In fact, between 1883 and

1991, 77 percent of southern bottomland hardwood forests

are estimated to have been lost due to timber harvesting,

conversion to agriculture, and other human uses (Gosselink

and Lee 1987, King and Keeland 1999, Mitsch and

Gosselink 2000, The Nature Conservancy 1992). The

combined efforts of the Wetlands Reserve Program (WRP)

and the Conservation Reserve Program (CRP) within the

U.S. Department of Agriculture (USDA), Natural Resources

Conservation Service (NRCS) and other Federal programs

have led to the planting and contracts for planting of about

126 000 ha of bottomland hardwoods in the Southern United

States since 1985 (Deavers 2000). We need sound methods

to evaluate these and various other wetland creation and

restoration efforts.

It has been shown that the natural regeneration of timbered

wetlands exhibits similar trends for soil organic matter

(SOM), total organic carbon (TOC), total Kjeldahl nitrogen

(TKN), and total phosphorus (TP) (Smith 1997). The Soil

Perturbation Index (SPI) evaluates how the different

successional stages compare to mature reference wetlands

from a biogeochemical standpoint. The SPI uses the means

of SOM, TOC, TKN, and TP data collectively for each

wetland to compare with the biogeochemical reference,

which is established from mature wetlands of the same type

and ecoregion. The SPI consists of data that were

transformed to percentage data by the following equation

using the parameter TP as an example:

[(u - c )/ u] x 100 = perturbation number,

where

u = TP mean value for the 0 (mature) successional stage

wetlands, and

c = TP mean for the cut wetland in question.

Percentages are calculated for all four parameters and then

plotted according to successional stage (fig. 1). Our

assumption is that the greater the change from the reference

conditions, the lower the ecological integrity of the system

(Maul and others 1999, Smith 1997). The SPI (Smith 1997)

indicates a shift in biogeochemical conditions after timber

harvest, with the greatest change from reference conditions

reached at approximately 8 to 9 years after alteration (fig. 1).

As calculated previously, this index predicts that it would

take 16 to 17 years for concentrations of the SPI

components (SOM, TOC, TKN, and TP) to return to

preharvest conditions (Maul and others 1999). We suggest

that the SPI be one component of an index of biotic integrity

for freshwater wetlands.

Our hypothesis was that if agriculturally based sites exhibit a

2d

 order polynomial curve as do logged and naturally

regenerated wetlands, one could use the SPI to evaluate

how the created and restored sites are developing. Thus, the

objectives of this study are to determine if reforested sites

previously in agriculture exhibit biogeochemical trends

similar to forested wetlands previously timber harvested

(Smith 1997), and, if so, to determine if the SPI is useful in

evaluating the restoration of created wetlands and restored

wetlands or both previously in agriculture. Soils from WRP,

CRP, and cottonwood sites within Mississippi, U.S.A., (fig. 2)

were analyzed to determine if reforested sites previously in

agriculture exhibit biogeochemical trends similar to those

exhibited by forested wetlands previously timber harvested
127

(as described by Smith 1997 and Smith and others 1996).

Because NRCS was interested in differences between WRP

and CRP, data are reported separately throughout our paper.

However, we were interested in overall SPI trends, so we

have shown WRP and CRP results on a single AG-SPI curve

(see fig. 6).

SITE DESCRIPTIONS

WRP sites with stands 3 and 5 years old and mature natural

forested wetlands with stands greater than 60 years old

were sampled in August and September of 1997 by Balducci

(1998) (table 1). Hereafter, the mature sites are referred to

as 0 successional stage because no creation or restoration

procedures were performed on those sites. These sites are

located in Leflore County, MS (fig. 2). In 1998, we focused on

sites reforested on agricultural land of various successional

stages ranging from 1 to 22 years since cultivation. To

encompass this range of successional stages, CRP and

cottonwood sites in Issaquena County, MS (fig. 2) were

sampled in August 1998 (table 1). The cottonwood sites

consist of two types: the term cottonwood refers to sites that

were planted on previous agricultural land and have not yet

been harvested; and coppice cottonwoods have been

through one harvest and resprouting cycle. The created sites

of table 1 include all WRP and CRP sites.

METHODS

Soil

Ten soil cores were randomly sampled within each site.

While in the field, cores were placed on ice and, once in the

lab, were frozen until processed. The top 5 cm were chosen

to represent the most recent influx of nutrients into the soil

(Smith 1997). The SOM was measured by loss on ignition at

550 oC for 5 hours (Craft and others 1991). The TOC was

analyzed by a Leco CN 2000 Carbon and Nitrogen Analyzer

(Leco Corporation 1996). The TKN and TP were measured

by a TRAACS 800 Autoanalyzer (Technicon Industrial

Systems Corporation 1987). Soil particle size analyses were

performed on a Horiba LA-910 Particle Size Analyzer

(Horiba Ltd. 1997).

Statistical Analyses

Individual parameter response—Each individual

parameter was analyzed separately using a one-way

analysis of variance (ANOVA) for each site. In the first

analysis, responses of SOM, TOC, TKN, and TP of the

cottonwood sites and mature sites (0 successional stage)

were observed. The treatment was successional stage

consisting of five groups: 0, 5, 8, 16, and 22 years (n = 3, 2,

2, 2, and 1, respectively). The second analysis compared

SOM, TOC, TKN, and TP among the WRP, CRP, and mature

sites. The treatment was successional stage consisting of

four groups: 0, 3, 5, and 9 years (n = 3, 3, 2, and 2,

respectively). The third analysis compared SOM, TOC, TKN,

and TP between 9-year coppice cottonwood sites and 9-year

CRP sites (n = 2). All variables were normally distributed

based on Shapiro-Wilk statistic for normality (SAS Institute,

Inc. 1988). All ANOVAs and Student-Newman-Kuels (SNK)

tests (p < 0.05) were performed with PC/SAS software (SAS

Institute, Inc. 1988).

Figure 1—The Soil Perturbation Index (SPI = percentage change of

SOM, TOC, TKN, and TP from uncut wetland means) from Maul and

others (1999) shows percent change from a biogeochemical

reference determined from mature wetlands prior to harvest (0 years

= highest ecological integrity exhibited in mature reference wetlands,

and 100 percent indicates greatest change from reference condition).

The SPI shows a decrease in biogeochemical function (55 percent)

that is greatest 8 to 9 years after timber harvesting. This index

predicts that it would take 16 to 17 years for the SOM, TOC, TKN,

and TP to return to preharvest conditions.

Figure 2—Map of Mississippi, U.S.A., showing Leflore

County (uppermost designated county) and Issaquena

County. Mature and WRP sites in Leflore County were

sampled by Balducci (1998) during 1997. CRP and all

cottonwood sites in Issaquena County were sampled during

1998. Map created by William A. Tedesco, Department of

Geology and Geological Engineering, The University of

Mississippi.
128

Soil Perturbation Index—The SPI (Maul and others 1999,

Smith 1997) uses the means of SOM, TOC, TKN, and TP

data for each wetland to calculate the percent change from

the biogeochemical reference determined from the mature

wetland soils. Because there are replicate wetlands per

successional stage (n) and four parameters analyzed per

wetland, 4 by (n) replicate points for each successional

stage are plotted. The cottonwood and mature sites were

used to develop a single agriculturally based SPI (AG-SPI)

for reforested sites previously in agriculture. As in Maul and

others (1999), the perturbation numbers were plotted for

each parameter (TP, SOM, TOC, and TKN) and for the

different successional stages. A 2d order polynomial

equation for AG-SPI [(Y = M0 + M1(x) + M2(x)) where M0 =

5.424, M1 = 10.660, M2 = -0.434, and x = successional

stage (years)] provided the best fit line for the index with

r = 0.77.

RESULTS

Individual Parameter Response

The majority of all the sites were classified as having soil

textural classes of silt or silt loam (table 2). The cottonwood

sites exhibited similar observable trends for all four

parameters (TP, SOM, TOC, and TKN). Concentrations of

soil parameters changed until the 8-year stage, then moved

back toward mature conditions (fig. 3). The SOM, TOC, TKN,

Table 1—Study sites, successional stage (years since alteration),

treatment (coppice, cottonwood, created, or natural), and soil nutrient

parameters

Site Stage Treatment SOM TOC TKN TP

Year - - - - - - - - - - - - mg/g - - - - - - - - - - - -

WRP3A 3 Createda 66.4 20.88 3.150 1.653

WRP3B 3 Createda 86.5 27.55 3.637 1.029

WRP3C 3 Createda 40.0 12.53 2.175 .475

5A 5 Cottonwoodb 88.8 29.01 3.380 1.206

5B 5 Cottonwoodb 52.1 14.40 1.863 .996

WRP5A 5 Createda 82.9 35.22 4.792 .745

WRP5B 5 Createda 30.6 5.24 1.213 .369

8A 8 Cottonwoodb 48.8 20.09 2.168 1.053

8B 8 Cottonwoodb 40.9 15.83 1.949 .845

C9A 9 Coppicec 101.6 36.71 4.972 1.339

C9B 9 Coppicec 95.3 32.79 4.216 1.171

CRPA 9 Createdd 76.2 20.22 2.535 .805

CRPB 9 Createdd 91.6 27.34 3.252 1.245

16A 16 Cottonwoodb 105.9 39.67 5.323 1.205

16B 16 Cottonwoodb 77.8 29.98 3.919 1.133

22A 22 Cottonwoodb 88.4 34.35 5.175 1.300

Mature > 60 Naturala 189.3 86.85 13.400 1.200

Mature > 60 Naturala 170.9 72.96 15.380 1.760

Mature > 60 Naturala 172.7 80.27 11.260 1.520

SOM = soil organic matter; TOC = total organic carbon; TKN = total Kjeldahl nitrogen;

TP = total phosphorus.

a Sites sampled by Balducci in August and September 1997.

b Cottonwood sites were planted on land previously used for agriculture and have not

been harvested (sampled August 1998).

c Coppice sites were allowed to sprout from trunks after harvesting (sampled August

1998).

d

 Conservation Reserve Program sites sampled in August 1998.
129

and TP trends exhibited in this study (fig. 4) are similar to

trends found in the timber-harvested wetland study sites of

Maul and others (1999). Coppice sites contained higher

concentrations of SOM, TOC, TKN, or TP than the created

sites (fig. 5).

Soil Perturbation Index

The AG-SPI combines data from the cottonwood and mature

sites (fig. 6). According to this study, reforested sites

previously in agriculture exhibit similar biogeochemical

trends as natural forested sites previously timber harvested.

The best-fit curve for both SPIs is a 2d order polynomial

curve. The AG-SPI indicates that reforested sites previously

in agriculture take longer to recover biogeochemically than

timber-harvested wetlands. The point indicating greatest

change from reference conditions for the AG-SPI is about 12

years whereas the original SPI indicates the greatest change

at 8 to 9 years after alteration. The AG-SPI predicts it would

take 24 years for SOM, TOC, TKN, and TP to return to

mature conditions compared to the 16 to 17 years suggested

by the original SPI. The AG-SPI biogeochemical levels

deviate approximately 70 percent from the mature sites

where the original SPI only changes about 55 percent from

the mature sites.

The WRP and CRP sites were plotted on the AG-SPI to

provide a biogeochemical comparison (fig. 6). The AG-SPI

indicates that WRP sites may exhibit a greater change from

mature biogeochemical conditions compared to other

wetlands of similar age and background. According to index

calculations run for the current study, the 9-year CRP sites

Table 2—Soil particle size analysis and textural class

data of Balducci (1998) and this study

Site Sand Silt Clay Textural class

- - - - - Percent - - - - - -

WRP3A 1.7 88.1 10.2 Silt

WRP3B 2.5 84.9 12.6 Silt loam

WRP3C 22.6 69.9 7.4 Silt loam

5A 14.8 80.1 5.1 Silt loam

5B 9.1 85.5 5.4 Silt

WRP5A 1.9 91.5 6.7 Silt

WRP5B 48.2 46.2 5.6 Sandy loam

8A 40.6 58.1 1.3 Silt loam

8B 39.0 59.6 1.4 Silt loam

C9A 7.1 87.6 5.3 Silt

C9B 8.0 87.1 4.9 Silt

CRPA 4.0 88.7 7.3 Silt

CRPB 8.1 83.7 8.2 Silt

16A 20.6 75.0 4.4 Silt loam

16B 29.7 67.0 3.3 Silt loam

22A 18.6 77.7 3.6 Silt loam

Mature 5.0 86.6 8.4 Silt

Mature 25.7 65.9 8.4 Silt loam

Mature 16.6 79.1 4.3 Silt

Figure 3—Trends of soil organic matter (SOM), total organic carbon (TOC), total Kjeldahl nitrogen (TKN),

and total phosphorus (TP) among cottonwood sites of different successional stages and mature sites. Error

bars represent one standard error. Data for mature sites are from Balducci (1998). Cottonwood sites were

sampled in August 1998.
130

Figure 4—Trends of soil organic matter (SOM), total organic carbon (TOC), total Kjeldahl nitrogen (TKN), and

total phosphorus (TP) among different successional stages of WRP (3 and 5 year), CRP (9 year) and mature

sites. Error bars represent one standard error. Mature and WRP sites were sampled in August and September

1997 (Balducci 1998). The CRP sites were sampled in August 1998.

Figure 5—Comparisons of soil organic matter (SOM), total organic carbon (TOC), total Kjeldahl nitrogen

(TKN), and total phosphorus (TP) between 9-year coppice and 9-year CRP sites. Error bars represent one

standard error. Sites sampled in August 1998.
131

are functioning in the same range as other sites

(cottonwood) previously in agriculture.

DISCUSSION

Individual Parameter Response

It is indicated in this study that reforested sites (cottonwood)

previously in agriculture exhibit similar biogeochemical

trends, although at lower levels, as wetlands allowed to

naturally regenerate following timber harvest. The SOM,

TOC, TKN, and TP concentrations decrease until the 8-year

stage then increase toward mature conditions. The pattern

observed in the current study is the same as that exhibited

by timber-harvested wetlands found in the Maul and others

(1999) study. Several factors promote such a change from

the biogeochemical parameters of mature wetlands and

subsequent rebound of soil nutrient concentrations. Young

aggrading stands have higher rates of nutrient uptake than

mature systems (Lockaby and Walbridge 1998). The young

successional stages produce less foliage and other organic

components contributing to organic matter than mature

systems. As the sites mature, the vegetation produces more

leaf litter, which upon recycling of nutrients, increases the

productivity of the wetland (Smith and others 1996).

Changes in the SOM concentrations or in rates of carbon

cycling may have important effects on nutrient cycling,

vegetative composition, and productivity (Trettin and others

1996). Thus, time is required in order for the system to reach

the biogeochemical conditions of mature reference wetlands.

Soil Perturbation Index

The SPI can aid as one component in evaluating created or

restored forested wetlands, or both, previously in agriculture.

The agriculturally based cottonwood sites produce an AGSPI very similar to that of the Maul and others (1999) SPI

developed from previously timber-harvested wetlands. The

recovery time of the AG-SPI is longer, and there is a greater

change in biogeochemical conditions than seen in the SPI

reported by Maul and others (1999). This longer recovery

time is expected due to the lack of organic matter

accumulation in the soil resulting from agricultural practices.

The disturbance of the soil and drier conditions of these

sites may contribute to the lower concentrations of SOM,

TOC, TKN, and TP retained within soils of the current study.

The WRP and CRP sites were plotted on the AG-SPI to

provide a biogeochemical comparison of how the created

wetlands are progressing. From a biogeochemical

standpoint, the 3- and 5-year WRP sites are functioning at

lower biogeochemical levels than other wetlands of the

same stage and background. Visual observation

(nonquantitative) indicated very few tree saplings were

present in any of the WRP sites. The 9-year CRP sites were

functioning within the range of other agriculturally based

sites (cottonwood) of the same stage. Visual observation

indicated trees were present and hardy in both CRP sites.

The SPI should be combined with other quantitative data

such as for vegetation and hydrology to provide a more

comprehensive evaluation of these created or restored

wetlands, or both. Further studies should focus on other

human alterations as well as natural disturbances. Wetlands

created for mitigation purposes might also be studied to

determine if the SPI is useful as one component in the

evaluation of these wetlands.

CONCLUSION

The SPI is a useful component and tool to evaluate created

or restored wetlands, or both, previously in agriculture. The

SPI is applicable to reforested wetland systems previously in

agriculture and can be one useful component of an index of

biotic integrity for wetland ecosystems. This study and that of

Maul and others (1999) indicate that wetlands previously

altered by human activity (agriculture and timber harvesting)

exhibit similar biogeochemical trends following alterations.

The SPI can be used to evaluate the biogeochemical trends

of created or restored wetlands, or both, (WRP and CRP)

compared to other wetlands of similar land use history, type,

and ecoregion.

ACKNOWLEDGMENTS

The authors thank Paul Rodrigue for valuable comments

and field assistance, Jonathan Maul for assistance with

statistical and graphical applications, and Tara Wildlife Inc.

for lodging. We are grateful for the use of equipment and

assistance from the USDA, ARS National Sedimentation

Laboratory in Oxford, MS. As with earlier studies, we are

grateful to Charlie Cooper and Al Mikell for thoughtful

discussions and suggestions. In addition, we appreciate the

helpful comments of Sammy King and Joy Zedler on an

earlier draft of this manuscript. This project was funded in

part by USDA, NRCS Wetland Science Institute (Agreement

68–7482–7–234 to The University of Mississippi,

Department of Biology, Marjorie M. Holland, Principal

Investigator).

Figure 6—The agriculturally based Soil Perturbation Index (AG-SPI)

shows percent change from a biogeochemical reference determined

from mature wetlands (0 years before alteration). The AG-SPI

shows a change in biogeochemical function (70 percent) that is

greatest 12 years after planting. This index predicts that it would

take 24 years for the SOM, TOC, TKN, and TP to return to mature

condition. The white circles represent mature and cottonwood sites

and the black circles represent WRP- and CRP-created sites. The

WRP and CRP sites were plotted for comparison. Mature and WRP

sites were sampled by Balducci (1998). Cottonwood and CRP sites

were sampled in August 1998.
132

REFERENCES

Balducci, J.A. 1998. Ecological integrity of wetland soils: testing of a

perturbation index in northern Mississippi wetlands. University,

MS: Department of Biology, The University of Mississippi. 63 p.

M.S. thesis.

Craft, C.B.; Seneca, E.D.; Broome, S.W. 1991. Loss on ignition and

Kjeldahl digestion for estimating organic carbon and total

nitrogen in estuarine marsh soils: calibration with dry combustion. Estuaries. 14: 175–179.

Deavers, L. 2000. Total acres enrolled in wetland reserve program.

USDA, Natural Resources Conservation Service. http://

www.wl.fb-net.org/wrp-acres.gif. [Accessed January 20, 2001.]

Gosselink, J.G.; Lee, L.C. 1987. Cumulative impact assessment in

bottomland hardwood forests. LSU–CEI–86–09. Center for

Wetland Resources. Baton Rouge, LA: Louisiana State University. 109 p.

Horiba, Ltd. 1997. Laser scattering particle size distribution analyzer

LA-910. Kyoto, Japan. 72 p.

King, S.L.; Keeland, B.D. 1999. Evaluation of reforestation in the

Lower Mississippi River Alluvial Valley. Restoration Ecology. 7(4):

1–12.

Leco Corporation. 1981. Leco instruction manual for CR12 carbon

analyzer No. 200-195. St. Joseph, MI: Leco Corporation.

Lockaby, B.G.; Walbridge, M.R. 1998. Biogeochemistry. In: Messina,

M.G.; Conner, W.H., eds. Southern forested wetlands. Boca

Raton, FL: CRC Press LLC: 149–172.

Maul, R.S.; Holland, M.M.; Mikell, A.T.; Cooper, C.M. 1999. Resilience

of forested wetlands located in the Southeastern United

States: demonstration of a soil perturbation index. Wetlands.

19: 288–295.

Mitsch, W.J.; Gosselink, J.G. 2000. Wetlands. New York: John Wiley

& Sons, Inc. 920 p.

SAS Institute, Inc. 1988. SAS/STAT user’s guide. Release 6. 3rd ed.

Cary, NC: SAS Institute Inc.

Smith, R.L. 1997. The resilience of bottomland hardwood wetlands

soils following timber harvest. University, MS: The University of

Mississippi. 99 p. M.S. thesis.

Smith, R.L.; Holland, M.M.; Cooper, C.M. 1996. The resilience of

bottomland hardwood wetland soils following timber harvest. In:

Heard, L.P.; Stewart, R.E., eds. Proceedings of a conference;

the Delta: connecting points of view for sustainable natural

resources. Memphis, TN: U.S. Department of Agriculture,

Natural Resources Conservation Service, Wildlife Habitat

Management Institute: 739–745.

Technicon Industrial Systems Corporation. 1987. TRAACS 800

method. Industrial method 787–86T; Elmford, NY: Braun and

Luebbe Analyzing Technologies, Inc.

The Nature Conservancy. 1992. The forested wetlands of the

Mississippi River: an ecosystem in crisis. Baton Rouge, LA: The

Nature Conservancy. 53 p.

Trettin, C.C.; Davidian, M.; Jurgenson, M.G.; Lea, R. 1996. Organic

matter decomposition following harvesting and site preparation

of a forested wetland. Soil Science Society of America Journal.

60: 1994–2003.
133

INTRODUCTION

In response to growing concern over agricultural pesticide

runoff effects on flora and fauna of receiving water bodies,

recent research has focused on potential best management

practices (BMPs) to minimize such effects. One such

suggestion is the use of constructed wetlands along field

perimeters to buffer receiving lakes, rivers, and streams

against potential effects of agricultural pesticides associated

with stormwater runoff. The U.S. Department of Agriculture

(USDA) established the new National Conservation Buffer

Initiative in 1997, which is aimed at installing 3.2 million km

of conservation buffers by the year 2002. In the Mississippi

Delta, areas containing wetlands have been drained for

agriculture and silviculture purposes. According to Mitsch

and Gosselink (1993), draining and filling of wetlands since

the mid-1800s has resulted in the loss of more than 50

percent of the Nation’s original wetlands. Perhaps by placing

constructed wetlands in areas where natural wetlands once

thrived, agriculture could regain the original wetland area

function of water-quality enhancement while minimizing

potential impacts to aquatic-receiving systems.

MATERIALS AND METHODS

A series of constructed wetland mesocosms specifically

designed to evaluate the fate of pesticides in wetlands was

used for this research (Rodgers and Dunn 1992). Darby

(1995) previously characterized wetland mesocosm

sediments. Four wetland mesocosms were chosen as test

cells, with one additional mesocosm serving as an

unamended control. Three remaining mesocosms were used

as water sources for the simulated storm event. Constructed

wetland mesocosms (59 to 73 m by 14 m) were amended on

two different occasions (one for the insecticide chlropyrifos,

the other for a mixture of the herbicides atrazine and

metolachlor). Following each pesticide amendment, a

simulated cropland runoff and rainfall event equal to three

volume additions was imposed on each wetland. Targeted

concentrations of chlorpyrifos were 73 μg/L, 147 μg/L, and

733 μg/L, whereas targeted concentrations for both atrazine

and metolachlor were 73 μg/L and 147 μg/L in addition to an

unamended control (0 μg/L). Water, sediment, and plant

samples were collected weekly for the duration of the

experiment (chlorpyrifos, 84 days; atrazine and metolachlor,

35 days). Samples were collected from sites longitudinally

distributed within each wetland and analyzed for the

respective pesticides using gas chromatography.

RESULTS AND DISCUSSION

Chlorpyrifos rapidly sorbed to sediment and plant material,

and the half-life in water, in this research, ranged from 5 to

13 days. Approximately 47 to 65 percent of measured

chlorpyrifos was within the first 30 to 36 m (from inflow) of

wetland mesocosms. Approximately 55 percent and 25

percent of measured chlorpyrifos mass were retained by

sediments and plant material, respectively. Conservative

mathematical models were used to derive adequate wetland

design parameters for mitigation of chlorpyrifos-associated

stormwater runoff. Recommended wetland travel distances

were 184 m for wetlands receiving 147 μg/L chlorpyrifos

runoff and 230 m for wetlands receiving either 73 or 733 μg/

L chlorpyrifos runoff. It is imperative to remember that these

calculations were made from conservative models.

Between 17 and 42 percent of the measured atrazine mass

was within the first 30 to 36 m (from inflow) of wetlands.

Atrazine concentrations were below analytical limits of

quantification (0.05 μg/L) in all sediment and plant samples

collected for this research. Atrazine aqueous half-lives

ranged from 16 to 48 days. According to conservative design

models, for atrazine concentrations of 73 μg/L, wetland

travel distances for effective mitigation ranged from 101 to

164 m. For wetlands receiving initial atrazine concentrations

of 147 μg/L, effective constructed wetland travel distances

ranged from 103 to 281 m.

Between 7 and 25 percent of measured metolachlor mass

was in the first 30 to 36 m (from inflow) of wetlands

immediately following application and simulated rainfall.

PESTICIDE MITIGATION CAPACITIES OF CONSTRUCTED WETLANDS

Matthew T. Moore, Charles M. Cooper, Sammie Smith, Jr., and John H. Rodgers, Jr.1

1 Moore, Cooperator, Cooper, Supervisory Ecologist and Research Leader, and Smith, Research Chemist, USDA-ARS National Sedimentation

Laboratory, P.O. Box 1157, Oxford, MS 38655; Rodgers, Professor and Director, Clemson Institute of Environmental Toxicology, 509

Westinghouse Road, Pendleton, SC 29670.

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability of

wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.

Abstract—This research focused on using constructed wetlands along field perimeters to buffer receiving water against

potential effects of pesticides associated with storm runoff. The current study incorporated wetland mesocosm sampling

following simulated runoff events using chlorpyrifos, atrazine, and metolachlor. Through this data collection and simple

model analysis, researchers conservatively predicted wetland buffer travel distances necessary to mitigate potential

pesticide effects on receiving systems, which ranged from 100 to 400 m for the pesticides. This research provides fundamental answers concerning the use of constructed wetlands for pesticide mitigation in agricultural watersheds.
134

Approximately 10 percent of measured metolachlor mass

was in plant samples. Based on models, adequate wetland

travel distances receiving 73 μg/L metolachlor would range

from 102 to 170 m. For those constructed wetlands receiving

initial concentrations of 147 μg/L metolachlor, wetland travel

distances necessary for effective mitigation would range

from 100 to 400 m.

Based on presented wetland designs, it may not be

economically feasible for a farmer to implement constructed

wetlands as a sole BMP if their fields were small, e.g., < 4

ha. Constructed wetlands are not panaceas for

environmental problems. Given proper situations, however,

they can greatly enhance water quality in a variety of

municipal, industrial, and agricultural settings. This research

has offered valuable data concerning effectiveness for using

constructed wetlands as buffers for pesticide-associated

stormwater runoff.

ACKNOWLEDGMENTS

Authors thank P.B. Rodrigue and the USDA NRCS Wetland

Science Institute for partial funding. Special thanks to G.M.

Huddleston; W.B. Gillespie, Jr.; R.W. Steinriede, Jr.; D.B.

Huggett; M. Baker; and D. Mathis for technical assistance

and advice.

REFERENCES

Darby, M.A. 1995. Modeling the fate of chlorpyrifos in constructed

wetlands. University, MS: The University of Mississippi. M.S.

thesis.

Mitsch, W.J.; Gosselink, J.G. 1993. Wetlands. New York:

VanNostrand Reinhold. 722 p.

Rodgers, J.H., Jr.; Dunn, A.W. 1992. Developing design guidelines

for constructed wetlands to remove pesticides from agricultural

runoff. Ecological Engineering. 1: 83–95.
135

INTRODUCTION

Artificial wetlands are now commonly used to improve water

quality; long-term operational information exists for many

full-scale projects (Kadlec and Knight 1986). Despite a

growing body of data relating artificial wetlands and waste

water, there have been few studies of the effects of natural

wetlands on waste water (see Kadlec and Knight 1986 for a

historical perspective). Those studies generally focused on

denitrification, which converts nitrates and ammonia to

nitrogen gas rather than on burial because, in most regions,

burial is too slow to remove significant nutrients. In subsiding

environments, however, burial might permanently store

significant amounts of nutrients. Nutrient burial rates have

been studied in natural wetlands in the Florida Everglades,

North Carolina estuarine marshes, and in southwestern

Louisiana (Craft and Richardson 1993, Craft and others

1993, Foret 1997). Nutrient storage rates in soil have rarely

been studied in natural wetlands receiving waste water and

never in wetlands of the Lake Pontchartrain Basin,

Louisiana. This is surprising given the large public concern

regarding eutrophication in the lake and the relatively rapid

subsidence of coastal Louisiana.

A lack of information on nutrient storage rates in wetlands

underlies public and professional controversy regarding the

benefits of water-quality regulations, wetland protection and

restoration activities, and the benefits of introducing nutrientrich water to estuarine marshes. Relationships among

nutrient sources and rates and nutrient sinks and rates need

to be better understood in wetland nutrient dynamics in

general, and in Lake Pontchartrain, in particular. We tested

the hypothesis that a natural wetland at the mouth of a

freshwater stream emptying into Lake Pontchartrain does

not affect nutrient inputs into the lake.

STUDY AREA

The study was conducted on the northern shore of Lake

Pontchartrain in a naturally occurring marsh that predates

the earliest aerial photographs. Lake Pontchartrain is a large

(the diameter is approximately 60 km) estuary in

southeastern Louisiana that receives freshwater from

numerous rivers and streams on its northern shore and sea

water from a tidal pass at its eastern end. The marsh studied

was located at the mouth of Salt Bayou, which carries

freshwater from the Pearl River to Lake Pontchartrain. The

marsh, which lies on a tributary (Salt Bayou) to Lake

Pontchartrain, was classified as brackish and intermediate

(Chabreck and Linscombe 1978). Construction of railways

and highways across Salt Bayou reduced freshwater input

from the Pearl River into the marsh in the mid-1950s. That

reduction in freshwater inflow is believed to have contributed

to subsequent conversion of marsh to shallow open water.

Since 1956, 2,485 ac of marsh have converted to shallow,

open water; 3,438 ac of marsh remained in 1990 (NRCS

1997).

A canal adjacent to Fritchie Marsh, the W-14 Canal, carries

urban runoff and tertiary-treated domestic sewage from the

community of Slidell to Lake Pontchartrain. Most water in the

W-14 Canal discharges directly into Lake Pontchartrain, but

some leaves the canal and flows through the Fritchie Marsh

before rejoining the W-14 Canal and then discharging into

Lake Pontchartrain. A Coastal Wetland Planning, Protection,

and Restoration Act (CWPPRA) project is scheduled for

construction in the Fritchie Marsh by the end of 1999. That

CWPPRA project will increase the amount of water flowing

from the W-14 Canal into the Fritchie Marsh. The intent of

that freshwater introduction is offset by the loss of freshwater

inputs from the Pearl River caused by the railways and

highways crossing Salt Bayou that were constructed earlier

this century. The Louisiana Department of Natural

Resources (DNR) will collect data that can be used to

estimate the effect of the W-14 Canal water on vegetation

and marsh loss in the Fritchie Marsh (Department of Natural

Resources 1996).

NUTRIENT STORAGE RATES IN A NATURAL MARSH RECEIVING WASTE WATER

J. A. Nyman1

1 Department of Biology, University of Louisiana at Lafayette, P.O. Box 42451, Lafayette, LA 70504-2451.

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability of

wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.

Abstract—Artificial wetlands are commonly used to improve water quality in rivers and the coastal zone. In most wetlands

associated with rivers, denitrification is probably the primary process that reduces nutrient loading. Where rivers meet

oceans, however, significant amounts of nutrients might be permanently buried in wetlands because of global sea-level rise

and regional subsidence. We determined nutrient storage rates in the marshes adjacent to a minor stream to test the

hypothesis that a natural wetland adjacent to Lake Pontchartrain (Louisiana) does not affect nutrient inputs to this estuary.

Data from soil cores indicated that marshes store 240 t of N and 11 t of P each year. These data demonstrate that wetlands

associated with rivers substantially modify nutrient transfer from terrestrial to aquatic habitats. Hydrologic data are being

collected that can be used to determine what portion of the nutrient load of the river is buried in the marsh, and what portion

of the nutrient load is discharged into the receiving basin.
136

METHODS

Marsh Soil

One data set was collected to quantify long-term nutrient

storage rates in Fritchie Marsh. Nutrient storage rates were

determined from the accretion rate of soil and the nutrient

content of soil in the marsh. This is possible because new

soil continually forms on the surface of estuarine marshes;

this process is called vertical accretion and is essential

because global sea-level rise and subsidence would

otherwise drown estuarine marshes (Mitsch and Gosselink

1986: 178–181). A pair of cores was collected from DNR

vegetative monitoring station 22; another pair of cores was

collected from DNR vegetative monitoring station 1. Station

1 was at the northern, upstream end of the study area;

station 22 was at the southern, downstream end. The cores

were returned to the lab, sectioned, weighed, ovendried,

weighed, and crushed. Bulk density of each section was

calculated. Vertical accretion since 1963 was determined

with the 137Cs dating technique (DeLaune and others 1978).

Activity of 137Cs was determined with an Ortec GMX series

Gamma-x-high purity, N-type germanium coaxial photon

detector system. Vertical accretion rates were compared to

regional estimates of subsidence to determine if there was a

vertical accretion deficit. Vertical accretion deficits are often

the mechanism by which sediment starvation and rapid

subsidence cause wetland loss in coastal Louisiana, e.g.,

Nyman and others 1993.

The gross material accumulation rate, i.e., the accumulation

of mineral sediments as well as organic matter, was

calculated for each core from the bulk density of soil

samples overlying the 1963 marsh surface. The

accumulation rate of mineral sediments and organic matter

was similarly calculated from the bulk density and ash

content of each section overlying the 1963 marsh surface.

The accumulation rate of nitrogen (N), carbon (C), and

phosphorus (P) was similarly calculated from the bulk

density and the nutrient content of soil overlying the 1963

marsh surface. Carbon and nitrogen content were

determined by ignition in a Carlo-Erba Elemental Analyzer

equipped with a thermoconductivity detector. Phosphorus

content was determined following acid digestion using

Parson’s colormetric analysis (Parsons 1984).

Water Quality

A second data set was collected to compare changes in

nutrient concentration in waters that empty directly into the

lake to changes in nutrient concentration in waters that flow

through the Fritchie Marsh before emptying into the lake.

Data were collected on two transects that originate at a

common point in the W-14 Canal and terminate at a common

point in Lake Pontchartrain. However, one transect runs

through the W-14 Canal, and the other runs through the

Fritchie Marsh. Selecting sites for collecting water samples

was more complicated than selecting sites for soil samples.

We had planned to collect water samples at DNR sites

simultaneously with DNR personnel. However, it was

impossible to sample along the W-14 Canal from the DNR

airboat because of numerous low bridges that cross the

canal. Alternate transportation was arranged after ULL

purchased a boat for research. Using a small boat rather than

an airboat reduced the area accessible for collecting water

samples and increased time required to collect samples

because shallow water areas cannot be crossed. Collecting

water samples was further delayed because the landowners

forbade sampling during waterfowl hunting season.

Both transects had 10 sampling stations. At each station, a

surface water sample was collected. Salinity and conductivity

were measured with a YSI model 6920 salinometer (Yellow

Springs, Ohio) simultaneously with the collection of each

water sample. Nutrient concentration was determined on the

unfiltered water samples. Total N and P content were

determined using Parson’s colormetric analysis (Parsons

1984). Changes in nutrient concentration resulting from

dilution with lake water were differentiated from nutrient

uptake with mixing diagrams, i.e., ionic ratios (Day and

others 1989: 81–85).

RESULTS

Marsh Soil

Average bulk density was greater at the northern site than at

the southern site. The greater bulk density at the northern

site incorrectly suggested that more material is introduced

into Fritchie Marsh from the W-14 Canal than from Lake

Pontchartrain. However, examination of density profiles with

depth suggests that recently formed, surface soils were

similar at the two sites. Thus, the northern and southern

sites currently appear to be storing more similar amounts of

material than the average bulk density suggests.

Accretion in Fritchie Marsh was slightly less than the

average for brackish marshes in the Mississippi River Deltaic

Plain (0.72 cm per year) (Nyman and others 1990). The

more rapid average accretion in the entire Mississippi River

Deltaic Plain than in the Fritchie Marsh results from the more

rapid subsidence rate elsewhere in the Mississippi River

Deltaic Plain (approximately 1.0 cm per year) (Penland and

others 1990).

Accretion rates within sites were almost identical, but the two

sites differed greatly. Accretion at the northern site was 40

percent slower than at the southern site. Accretion at the

northern site is similar to estimates of subsidence on the

northern shore of Lake Pontchartrain, which is estimated at

0.45 cm per year (Penland and Ramsey 1990). However, it is

difficult to reconcile the rapid accretion at the southern site

with the slow subsidence reported by Penland and Ramsey

(1990). It is also unusual for accretion rates to vary so much

within such a small area, e.g., Nyman and others 1990,

Nyman and others 1993. The large difference between

northern and southern sites suggests that a shallow, active

fault runs through Fritchie Marsh. Such faults are common in

coastal Louisiana in general and around Lake Pontchartrain.

Lopez (1991) examined seismic data and reinterpreted the

location of the Baton Rouge/Denham Springs fault system

and concluded that the fault lay farther south than previously

believed and positioned it such that it would bisect the

Fritchie Marsh. Our findings support Lopez’s (1991)

conclusions regarding the position of that fault system.

Lopez (1991) also concluded that the fault system was

active and responsible for a 6-inch offset on the State

Highway Bridge 11 crossing eastern Lake Pontchartrain. The

tremendous difference in accretion rates between the
137

northern and southern sites that we observed support

Lopez’s (1991) conclusion that the fault system is active,

although different faults within the fault system would be

required to produce offset in accretion in the Fritchie Marsh

and the offset in the highway bridge. Whereas this active

fault may contribute greatly to wetland loss in the Fritchie

Marsh, it also increases the potential for burial of nutrients in

the marsh.

Gross material accumulation was 20 percent slower at the

northern site than at the southern site. The faster

accumulation rate at the southern site results from the more

rapid vertical accretion induced at the southern site by the

fault. Extrapolated to the entire area, it appears that the

1,040-ac Fritchie Marsh restoration site stores slightly over

13,660 t of material annually. Some of this material is

organic carbon that is produced in the marsh, but the

associated nutrients and sediments otherwise would be

discharged into Lake Pontchartrain.

The material being stored in wetland soil at the Fritchie

Marsh includes mineral sediments, organic matter, and the

ecologically important elements C, N, and P. A significant

amount of the nutrients N and P is being stored in the marsh

soils, but the amount stored has decreased 48 percent since

the 1950s because of the conversion of wetlands to shallow

open water areas, which are assumed to be stable rather

than accreting.

The N:P ratio of soil at the northern site was slightly higher

than that at the southern site. Higher ratios indicate a greater

potential for P availability to limit plant growth at the northern

site, but the small difference between the sites may not be

ecologically meaningful. These ratios appeared typical; they

were similar to those reported for unmanaged, S. patens

dominated marsh at Rockefeller Refuge (Foret 1997).

Water Quality Data

Nutrients were more concentrated in water in the W-14

Canal than in Lake Pontchartrain water. This situation is

typical of estuaries in general (Liss 1976). For example, N

concentrations in the Mississippi River generally average 3

to 4 parts per million (ppm) but drop rapidly to approximately

1.0 ppm soon after entering estuarine marshes (Lane and

others 1995). Phosphorus in the Mississippi River also drops

rapidly soon after entering estuarine marshes from

approximately 0.4 ppm P in the river to approximately 0.2

ppm P after entering estuarine marshes. Nutrient

concentrations in the W-14 Canal were less than those

reported for the Mississippi River (Antweiler and others

1995, Lane and others 1999) and did not exceed 1.5 ppm N

(108 μmole N/l) or 0.5 ppm P (16 μmole P/l).

A simple comparison of nutrient concentrations between

water in the W-14 Canal and the marsh incorrectly suggests

that the marsh was removing nutrients when the water

samples were collected. Proper evaluation of the waterquality data indicates that the low-nutrient water in the

marsh was actually low-nutrient, high-salinity water

introduced from Lake Pontchartrain at an earlier date and,

subsequently, stored in the marsh. Thus, the low-nutrient,

high-salinity water in the marsh interior indicates at least

periodic hydrologic isolation of the marsh interior. The

isolation apparently results from a lack of inflow from the W14 Canal at the northern (upstream) end of the marsh, and

possibly from high-water levels in the W-14 Canal that inhibit

drainage of the marsh at the southern (downstream) end.

The proposed restoration project should introduce more

water through the marsh and thereby reduce the hydrologic

isolation that the marsh currently experiences. It could not

be determined if the failure to detect nutrient uptake with the

water-quality data resulted from a true lack of uptake or from

an inability of the technique to detect real differences.

Water in the W-14 Canal had a very low N:P ratio relative to

N:P ratio in the marsh interior. The difference in N:P ratio

resulted from 7-fold lower P concentrations rather than

greater N concentrations. The N:P ratio of the W-14 Canal

was similar to ratios in P-rich waters entering the Florida

Everglades; the N:P ratio of the marsh interior was similar to

ratios in unimpacted portions of the Florida Everglades

(Reddy and others 1999). The N:P ratio in water samples

suggests that P availability limits algae production in the

marsh interior but not the W-14 Canal. Increasing input of Prich water to the marsh should increase algae production,

which might increase the formation of recalcitrant, organic-P

compounds.

CONCLUSIONS

Data from four cores indicated that the marshes at the

Fritchie Marsh store significant amounts of N and P each

year. Collectively, the marshes remove 240 t of N and 10.9 t

of P from the Lake Pontchartrain estuary each year. While

significant, these rates are only 42 percent of the removal

rates that existed before extensive wetland loss that

occurred since 1956.

Nutrient storage in wetland soils was faster at the southern

end of the marshes despite nutrient concentrations being

lower because soil formation there is more rapid. More rapid

soil formation at the southern end of the marshes appears to

result from an active fault system previously reported in the

area. Additional data are needed to characterize nutrient

storage given the great difference between the northern and

southern ends of the study area.

Water-quality data indicated limited water exchange

between the marsh interior and the adjacent water channels.

Water-quality data failed to indicate nutrient uptake when

mixing diagrams were used, but N:P ratios in water samples

indicated that algae production in the marsh interior was Plimited when samples were collected.

The imminent wetland restoration project should increase

nutrient storage rates in marsh soil and may increase the

production of recalcitrant, organic forms of P that can be

stored in bottom sediments of shallow, open-water areas

common in the marsh interior. The imminent wetland

restoration project should also reduce the hydrologic

isolation of the marsh interior and may thereby reduce plant

stress, increase plant production, and increase nutrient

burial in soil.
138

REFERENCES

Antweiler, R.C.; Goolsby, D.A.; Taylor, H.E. 1995. Nutrients in the

Mississippi River. In: Meade, R.H., ed. Contaminants in the

Mississippi River, 1987–92. U.S. Geol. Sur. Circ. 1133. Denver:

U.S. Geological Survey: 73–86.

Chabreck, R.H.; Linscombe, G. 1978. Vegetative type map of the

Louisiana coastal marshes. Baton Rouge, LA: Louisiana

Department of Wildlife and Fisheries.

Craft, C.B.; Richardson, C.J. 1993. Peat accumulation and N, P, and

organic C accumulation in nutrient-enriched and natural

everglades peatlands. Ecological Applications. 3: 446–458.

Craft, C.B.; Seneca, E.D.; Broome, S.W. 1993. Vertical accretion in

microtidal regularly and irregularly flooded estuarine marshes.

Estuarine, Coastal and Shelf Science. 37: 371–386.

Day, J.W.; Hall, C.A.S.; Kemp, W.M.; Yanez-Arancibia, A. 1989.

Estuarine ecology. New York: John Wiley. 576 p.

DeLaune, R.D.; Patrick, W.H., Jr.; Buresh, R.J. 1978. Sedimentation

rates determined by 137Cs dating in a rapidly accreting salt

marsh. Nature. 275: 532–533.

Department of Natural Resources. 1996. Monitoring plan. Proj. PO–

06; Fritchie Marsh. Baton Rouge, LA: Louisiana Department of

Natural Resources, Coastal Restoration Division.

Foret, J.D. 1997. Accretion, sedimentation, and nutrient accumulation rates as influenced by manipulations in marsh hydrology in

the Chenier Plain, Louisiana. University of Southwestern

Louisiana. 56 p. M.S. thesis.

Kadlec, R.H.; Knight, R.L. 1996. Treatment wetlands. Boca Raton,

FL: CRC Press, Inc. 893 p.

Lane, R.R.; Day, J.W., Jr.; Thibodeaux, B. 1999. Water quality

analyses of a freshwater diversion at Caernarvon, Louisiana.

Estuaries. 22: 327–336.

Liss, P.S. 1976. Conservative and non-conservative behaviour of

dissolved constituents during estuarine mixing. In: Burton, J.D.;

Liss, P.S., eds. Estuarine chemistry. New York: Academic Press:

93–130.

Lopez, J.A. 1991. Origin of Lake Pontchartrain and the 1987 Irish

Bayou earthquake. In: Coastal depositional systems of the Gulf

of Mexico: quaternary framework and environmental issues. Gulf

Coast Section Society of Economic Paleontologists and

Mineralogists Foundation, 12th annual research conference;

Austin, TX: Earth Enterprises: 103–110.

Mitsch, W.J.; Gosselink, J.G. 1986. Wetlands. Van Nostrand

Reinhold Co. 539 p.

Natural Resources Conservation Service. 1997. Project plan and

environmental assessment, Fritchie Marsh hydrologic restoration; PO–6; St. Tammany Parish. Alexandria, LA: U.S. Department of Agriculture, Natural Resources Conservation Service.

Nyman, J.A.; DeLaune, R.D.; Patrick, W.H., Jr. 1990. Wetland soil

formation in the rapidly subsiding Mississippi River Deltaic Plain:

mineral and organic matter relationships. Estuarine, Coastal and

Shelf Science. 31: 57–69.

Nyman, J.A.; DeLaune, R.D.; Patrick, W.H., Jr. 1993. Relationship

between vegetation and soil formation in a rapidly submerging

coastal marsh. Marine Ecology Progress Series. 96: 269–279.

Parsons, T.R. 1984. A manual of chemical and biological methods

for seawater analysis. New York: Pergamon Press. 173 p.

Penland, S.; Ramsey, K.E. 1990. Relative sea-level rise in Louisiana

and the Gulf of Mexico: 1908–1988. Journal of Coastal Research. 6: 323–342.

Reddy, K.R.; White, J.R.; Wright, A.; Chua, T. 1999. Influence of

phosphorus loading on microbial processes in the soil and water

column of wetlands. In: Reddy, K.R.; O’Conner, G.A.; Schelske,

C.L., eds. Phoshporus biogeochemistry in subtropical ecosystems. Boca Raton, FL: CRC Press: 249–273.
139

EFFECTS OF SOIL OXIDATION-REDUCTION CONDITIONS ON

INTERNAL OXYGEN TRANSPORT, ROOT AERATION, AND GROWTH

OF WETLAND PLANTS

S.R. Pezeshki and R.D. DeLaune1

Abstract—Characterization of hydric soils and the relationship between soil oxidation-reduction processes and wetland

plant distribution are critical to the identification and delineation of wetlands and to our understanding of soil processes and

plant functioning in wetland ecosystems. However, the information on the relationship between flood response of wetland

plants and reducing soil conditions is limited. We have examined the influence of intensity and capacity of soil reduction on

internal oxygen transport, rhizosphere oxygenation, nutrient uptake, root and shoot growth, and survival of several wetland

species. Whereas the study species displayed a wide range of responses, intense soil reduction below -200 mV adversely

affected growth and biomass accumulation in the majority of these species. It is clear that high oxygen demand in soil

resulting from intense reduction influences oxygen transport and release to the rhizosphere. In addition, root elongation and

shoot growth are profoundly influenced by the intensity and capacity of soil reduction.

1

 Pezeshki, Professor of Biology, Department of Biology, University of Memphis, Memphis 38152, TN; DeLaune, Professor, Wetland Biogeochemistry Institute, Louisiana State University, Baton Rouge, LA 70803.

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability of

wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.

INTRODUCTION

Numerous reviews and book chapters on plant responses to

soil waterlogging are available that provide detailed literature

synthesis on this topic (Armstrong and others 1994; Drew

1990, 1997; Hook and Crawford 1978; Jackson and others

1991; Kozlowski 1984a, 1984b, 1997; Pezeshki 1994;

Vartapetian and Jackson 1997). Despite this wealth of

literature, close examinations reveal that in most studies the

status of soil oxidation-reduction conditions has not been

reported. Although some researchers have reported oxygen

concentrations in the root medium, such measure in wetland

soil does not provide adequate information to allow

evaluation of the intensity of soil reduction (DeLaune and

Pezeshki 1991). This point is important because in wetland

systems most plants are well adapted to endure soil-oxygen

deficiency but may differ in ability to withstand certain levels

of intense soil-reducing conditions.

In a typical flooded wetland soil, plants respond to the soil

physicochemical changes. These responses may lead to a

wide range of plant-stress symptoms.

Although various plant responses to flooded soil conditions

have been addressed in numerous publications, little

information can be found on the relationship between

wetland plant functions and the two aspects of soil redox

potential—the intensity and the capacity of reduction. The

reduction of the inorganic redox systems in a flooded soil

may be characterized in intensity or capacity terms. The

intensity factor determines the relative ease of reduction

whereas the capacity factor refers to the amount of the

redox system undergoing reduction (DeLaune and Pezeshki

1991). From a plant physiological-ecology standpoint, there

are many uses of interpretations of redox processes in soils;

one example is that the knowledge of the soil redox potential

represents an indication of the oxidation-reduction status of

various soil compounds. For example, a redox potential of

zero mV indicates that oxygen and nitrate are not likely to be

present and that the bioreducible iron and manganese

compounds are in a reduced state. At this same potential,

however, sulfate is stable in the soil with no sulfide being

formed, which is toxic to plants. A redox potential of +400

mV indicates that oxygen may be present even though there

may be excess water in the soil (DeLaune and Pezeshki

1991). Thus, the primary objective of the present paper is to

summarize and assess evidences on the significance of the

intensity and capacity of reduction in soils to wetland plant

functioning. We will emphasize the relationships between

soil flooding, reduced soil conditions (low soil redox

potential, Eh), the components of the soil oxidation-reduction

system, namely the intensity and the capacity of reduction,

and their influence on internal oxygen transport, rhizosphere

oxygenation, root and shoot growth, and survival of wetland

species.

Soil Oxidation-Reduction Potential (Eh)

Soil flooding initiates a chain of reactions leading to reduced

soil conditions (low soil redox potential, Eh). These reactions

include physical, chemical, and biological processes that

have significant implications for wetland plant functioning,

survival, and productivity (see Gambrell and others 1991,

Gambrell and Patrick 1978, Ponnamperuma 1984). Physical

processes include restriction of soil-atmospheric gas

exchange and depletion of soil oxygen needed for root

respiration. Once flooded, the limited supply of oxygen in

floodwater is depleted by roots, soil microorganisms, and

soil reductants rapidly (Ponnamperuma 1972). The depletion

of oxygen results in a series of chemical changes in soil

including accumulation of CO2, methane, N2, and H2

(Ponnamperuma 1984). The processes that follow include

denitrification, reduction of iron, manganese and sulfate, and
140

changing soil pH and Eh (Gambrell and others 1991). In a

typical series of reduction, NO3

-

 is reduced to NO2

-

, followed

by reductions of Mn+4 to Mn+2, Fe+3 to Fe+2, SO4

2-

 to S2- and

accumulations of acetic and butyric acids produced by

microbial metabolism (Gambrell and Patrick 1978,

Ponnamperuma 1984). Thus, soil redox potential decreases

(become more negative) in response to flooding (Gambrell

and Patrick 1978; Patrick and DeLaune 1972, 1977;

Ponnamperuma 1984). Aerated soils have characteristic

redox potentials in the range of +400 to +700 mV, whereas

waterlogged soils may exhibit redox potentials as low as -

300 mV.

In wetland soils, plants are faced with not only the lack of

oxygen but also a substantial demand for oxygen in root

medium (DeLaune and others 1990, van Wijck and others

1992). Such conditions create the potential for root oxygen

loss to soil; thus, additional root stress (Brix and Sorrell

1996, DeLaune and others 1990).

Quantifying Soil Redox Conditions

Many of the terms used in the literature to define flooded soil

conditions are difficult to quantify. For example, “flooded”,

“saturated”, or “waterlogged” are frequently used to describe

low oxygen supply conditions in soils. However, these terms

hardly define root-zone conditions (DeLaune and Pezeshki

1991). In addition, due to the absence of oxygen in most

“waterlogged” soils, methods used for the measurement of

oxygen content and oxygen diffusion rate in well-drained

soils cannot be used effectively (Gambrell and others 1991).

In contrast to the mostly nonquantifiable terms mentioned

above, soil Eh is a useful term because it can be measured

easily in the laboratory and in the field. The electrodes are

relatively easy to build and handle under most conditions

(for details see Faulkner and others 1989; Patrick and

DeLaune 1972, 1977). Furthermore, quantifying soil Eh is

particularly advantageous in periodically flooded soils

because the range of Eh is much wider, ranging between

approximately -300 to +700 mV, than either aerated or

permanently waterlogged soils (DeLaune and Pezeshki

1991). Thus, soil Eh measurement is perhaps the best

available quantifying tool of defining soil chemical status in

wetland systems (Patrick and DeLaune 1977).

Intensity and Capacity of Reduction

The reduction of the inorganic redox systems including

oxygen in hydric soils following flooding can be described in

“intensity” and “capacity” terms. The intensity factor

determines the relative ease of the reduction, whereas the

capacity factor denotes the amount of the redox system

undergoing reduction, e.g., oxygen consumption at root

interface. It is represented by the free energy of the

reduction, or, more commonly, by the equivalent

electromotive force (EMF) of the reactions (Patrick and

others 1986). In natural systems such as soils where there is

biological activity and where many redox systems function,

the oxidation-reduction or redox potential is ordinarily used

to denote the intensity of reduction. The capacity factor of a

redox system probably can be best described in terms of its

oxygen equivalent (Kludze and DeLaune 1995, Reddy and

others 1980). The capacity factor of the various redox

systems can vary from one soil to another.

Proper evaluation of wetland plant responses to soil flooding

requires evaluation of both intensity and capacity of soil

reduction because these two components influence oxygen

demand (Kludze and DeLaune 1995). DeLaune and others

(1990) were the first to demonstrate that oxygen demand

(capacity of reduction) in conjunction with intensity of

reduction (as determined from soil Eh measurements) in the

root medium was important to predicting wetland plant

functioning. Furthermore, they noted that using oxygendepleted solutions by nitrogen gas introduction did not

represent a high root-oxygen demand environment; thus,

they were a poor analogue of wetland soils. DeLaune and

others (1990), Kludze and others (1993), and Sorrel and

others (1993), using titanium citrate solution to create a high

oxygen-demand root environment, reported that root oxygen

transport and release were affected by such conditions in

several wetland species. However, such a solution, while a

significant improvement over de-oxygenated solution, at best

mimics wet soil conditions (Kludze and DeLaune 1995); but,

it does not represent other important characteristics of wet

soils including the soil’s capacity for phytotoxin production

that has a significant effect on most wetland species. There

is a critical need for development of reliable methods for

quantifying capacity of reduction in soils to complement the

more easily measured reduction intensity (Eh).

Plant Responses to Reducing Soil Conditions:

The Internal Oxygen Transport System

In wetland plants, usually an extensive oxygen transport

system of aerenchyma tissue facilitates oxygen diffusion

from aerial parts to the roots. Such a system may exist in

roots, stems, and leaves but is found primarily in roots

(Armstrong and others 1994). This system allows a plant to

transport the needed oxygen to the roots for maintaining

aerobic respiration and to oxidize reducing compounds in

the rhizosphere. In addition, the internal system of large gas

spaces also reduces internal volume of respiring tissues and

oxygen consumption, thus, enhancing the potential for

oxygen reaching the distant underground portions of the

plant (Armstrong and others 1994, 1996). Due to such

advantages, the oxygen transport system has been

considered as a major mechanism critical to a plant’s ability

to cope with soil anaerobiosis (Armstrong and others 1991,

1994, 1996; Drew 1990, 1992, 1997; Hook and Crawford

1978; Kozlowski 1982, 1984a, 1984b, 1997; Pezeshki 1994,

Teal and Kanwisher 1966).

Root oxygenation helps plants overcome intense

anaerobiosis (Armstrong and others 1994, 1996) and has

important ecological implications in wetlands. For instance,

the vigor and productivity of Spartina alterniflora were found

to be positively correlated with substrate redox potentials

because of the interaction with root aeration (DeLaune and

others 1983). Two related factors contribute to limit plant

growth under highly reduced soil conditions. First, the low Eh

levels represent an oxygen-deficient system. Second, soil

phytotoxins, the byproducts of low soil Eh, may continue to

accumulate to concentrations that the roots’ oxidizing power

can no longer adequately ameliorate them (Mendelssohn

and others 1981, Ponnamperuma 1984, Teal and Kanwisher

1966). Under such conditions, the roots must rely more

heavily on anaerobic respiration or transport sufficient
141

oxygen to roots to maintain aerobic respiration, lessening its

capacity to oxidize the rhizosphere.

Internal plant air space results from aerenchyma tissue

formation that is enhanced in many wetland species in

response to flooding and is critical to wetland plant

functioning in flooded soils (Drew and others 1985,

Kozlowski 1997, Schat 1984). The effectiveness of the gas

transport is primarily dependent on two factors: first, the

resistance to diffusion that is proportional to root length and

inversely proportional to root porosity; second, the oxygen

demand along the diffusion path resulting from respiratory

needs as well as oxygen leakage from the roots into the

rhizosphere (Armstrong 1979, Luxmoore and others 1972). It

is known that oxygen demands of roots and rhizosphere are

competitive because, in flooded soils, these systems

compete for the plant pool of oxygen simultaneously

(Armstrong and Beckett 1987; Armstrong and others 1991,

1994). As soil reduction continues, there is a progressively

greater demand imposed upon roots for oxygen (Delaune

and others 1990).

Soil Oxidation-Reduction, Plant Oxygen Transport,

Functioning, and Growth

There is a limited body of data on the relationship between

functional aspects of gas transport within plants and soil

oxidation-reduction conditions. In a few studies that

evaluated the relationship between plant responses and the

intensity of soil reduction, it became evident that intense soil

reduction (low soil Eh) promoted oxygen loss from root to

the rhizosphere (fig. 1). For instance, in some wetland

species, high correlation (r = 0.96) was found between radial

oxygen loss (ROL) from roots and soil Eh intensity; i.e., there

was an increasingly higher oxygen loss rate as soil Eh

became more reduced (Kludze and DeLaune 1995). In

addition, low soil Eh led to decreased leaf carbon

assimilation (fig. 2) and substantial inhibition of root

elongation (fig. 3).

In S. patens, a dominant U.S. gulf coastal marsh species,

root porosity increased as soil Eh decreased, resulting in

root porosity of 22 percent in plants grown at +200 mV

whereas porosity was 45 percent in plants grown at –300

mV. Also, ROL was significantly greater for plants in –300

mV Eh treatment as compared to +200 mV Eh (Kludze and

DeLaune 1994). Other studies have shown similar

responses for root porosity-soil Eh intensity relationship in

other wetland plants including swamp and bottomland

woody species (Kludze and others 1994, Pezeshki and

Anderson 1997). In contrast, Brix and Sorrell (1996)

reported that root porosity in two wetland species, Phalaris

arundinacea and Glyceria maxima, did not change in

response to reducing Eh in root medium to as low as –250

mV. However, their experiment lasted only 12 days, of which

the Eh in the root medium was at –250 mV for the last 7

days; thus, plants may not have had adequate time for

response. Alternatively, the two species may have already

had full-scale porosity development prior to treatment

initiation. However, this is an unlikely scenario because

porosity ranged from 5.3 to 11.6 percent, well below the

range reported for many wetland plants. The range of

porosity in many wetland species is much higher, particularly

if grown under low soil Eh conditions (Kludze and DeLaune

1994, Kludze and others 1994). In a study conducted on

Taxodium distichum, significant increase in root porosity and

ROL was noted at Eh intensity of –240 to –260 mV in root

medium (Kludze and others 1994). ROL increased from 12.7

in control to 42.3 mmol O2

 per gram per day. Similarly, root

air space increased from 13.3 to 41.4 percent in response to

the intensity of reduction. In Oryza sativa (rice), ROL

increased in response to a drop in soil Eh concomitant with

root porosity that increased from 26.8 to 35 percent when Eh

dropped from +200 to –300 mV (Kludze and others 1993).

Despite the reported increase in aerenchyma tissue

formation in many wetland species and, thus, the increase in

porosity in response to reducing soil conditions, such

increase may not be sufficient to satisfy the root respiratory

need for oxygen, perhaps due to the greater ROL in

response to high intensity of reduction. Pezeshki and others

Figure 1—Root porosity (POR) and radial oxygen loss (ROL) in

Spartina patens grown under various soil redox intensity for 50 days.

Bars represent SE. (Redrawn from Kludze 1994.)
142

(1991, 1993) concluded that despite a substantial

enhancement of aerenchyma tissue formation in Spartina

patens, alcohol dehydrogenase activity continued to be

higher in flooded than control plants indicating continued

oxygen stress in the roots of flooded plants. In addition, the

increase in ROL reported under intense soil Eh may explain

the reported reductions in root growth of several wetland

species under low soil Eh conditions. For instance, in

Spartina patens root and shoot dry weights decreased by 40

percent and 25 percent as soil Eh dropped from +200 mV to

–300 mV, respectively. Results clearly indicated the

influence of soil Eh intensity on growth of this marsh

species. It also was demonstrated that roots were more

sensitive to Eh intensity than shoots (Kludze and DeLaune

1994). Pezeshki and DeLaune (1990) reported cessation of

root growth in S. patens at soil Eh below –100 mV (fig. 3). In

addition, Pezeshki and others (1991) noted smaller root

systems in S. patens under reducing conditions and

concluded that such reduction in sink size may, in part, be

responsible for a negative feed-back inhibition of

photosynthesis resulting in a reduction in productivity of this

species. DeLaune and others (1990) studied plant

responses to the intensity of soil reduction using a titaniumcitrate solution. They demonstrated that the intensity of the

reduction in growth medium and the resulting demand for

oxygen in the root zone exerted significant influence on plant

physiological functioning.

The redox capacity factor is also important although much

less is know about its effects on wetland plants than is

known about the effects of the intensity factor. In fact, two

different soils with the same level of intensity of reduction

may differ substantially in the capacity for reduction. Soil

reduction capacity can be determined using measurements

of soil respiration CO2 and calculating oxygen equivalent by

stoichiometry (Kludze and Delaune 1995). Levels of soil

redox capacity may be created or manipulated, or both, by

providing extra carbon and energy source (organic matter) to

the soil while maintaining the same redox intensity level. In

an experimental set-up, reduction capacity may be

controlled by adding different amounts of granular D-glucose

to the root medium, which is also maintained under preset

reducing conditions (Eh< +350 mV) (Kludze 1994).

As was the case for the intensity of Eh reduction, differences

in Eh capacity among wetland soils may influence many

plant functions (figs. 4, 5) including oxygen transport,

rhizosphere oxygenation, photosynthetic rates, and, thus,

many aspects of plant functioning (Kludze and DeLaune

1995). Studies showed that increased Eh capacity under a

constant Eh intensity of –200 mV did not have any

significant effect on root porosity in S. patens, but oxygen

release was increased in response to the increasing Eh

capacity (Kludze and DeLaune 1995). However, the authors

reported that there was a threshold of Eh capacity beyond

which oxygen release remained constant or decreased, or

both, in this species. The response was attributed to the

potential effects of several factors such as soil phytotoxins

as well as plant physiological responses including stomatal

closure. However, the reasons for such a response remain

unknown. Plant carbon fixation, root, and shoot growth were

significantly inhibited in S. patens under increasing soilreduction capacity. Root and shoot dry weights decreased

Figure 2—Carbon assimilation in Spartina patens grown under

various soil reduction intensities. Data were collected at 17 days

after growth in incubated soils. Values followed by the same letter

are not significantly different at the 0.05 level. (Redrawn from Kludze

1994.)

Figure 3—Root elongation in Spartina patens grown under various

soil reduction intensities (Eh, mV). Data were collected on plants

grown in different rhizotrons in which soil Eh was manipulated using

various air/nitrogen gas mixtures. (Redrawn from Pezeshki and

DeLaune 1990.)
143

by 70 and 37 percent in high reduction capacity conditions

compared to control plants, respectively (Kludze and

DeLaune 1995).

In addition to the effects on growth, limited data suggest that

both the intensity and capacity of reduction may govern

nutrient uptake in wetland plants. In a study of seedlings of

two bottomland woody species grown in soil suspension

maintained at three Eh levels, +560, +340, and +175 mV,

fertilizer 15N uptake decreased with decreasing soil redox

potential, a response to the intensity of reduction (DeLaune

and others 1998). Phosphorus uptake by Typha

domingensis was inhibited in response to decreases in soil

redox potential (intensity of reduction) in rooting medium.

Increasing capacity of reduction (using a titanium citrate

solution) resulted in a further decrease in phosphorus

uptake (DeLaune and others 1999), further confirming the

effects of reduction capacity on nutrient uptake.

CONCLUDING REMARKS

Based on the limited data discussed above, both intensity

and capacity of reduction appear to influence plant

functioning in wetland ecosystems. In wetland soils, plants

are faced with a substantial demand for oxygen in the

rhizosphere and the potential for loss of oxygen to the soil

resulting in additional root stress. As soil reduction continues

and intensifies, there is a progressively greater demand

imposed upon roots for oxygen, and, thus, a greater

potential for loss of oxygen to the rhizosphere. The severity

of oxygen loss and the effects of reduction intensity and

capacity on plant functioning appear to be broad across

wetland species. The need for additional data on various

aspects of plant functioning and growth in wetland

ecosystems in response to soil redox conditions, especially

capacity of reduction, is clear. Redox potential

measurements collected in soils using platinum electrodes

provide information only on intensity of reduction. In studying

plant response to a reducing environment, methods must be

developed to evaluate the influence of capacity of soil

reduction in addition to the intensity of soil reduction on

wetland plant growth and functioning.

ACKNOWLEDGMENTS

The authors are thankful to Drs. I. Devai and A. Jugsujinda

for helpful comments made on an earlier version of this

manuscript.

REFERENCES

Armstrong, J.; Armstrong, W.; Van Der Putten, W.H. 1996.

Phragmites die-back: bud and root death, blockage within the

aeration and vascular systems and the possible role of

phytotoxins. New Phytologist. 133: 399–414.

Armstrong, W. 1979. Aeration in higher plants. Advances in

Botanical Research. 7: 225–332.

Armstrong, W.; Beckett, P.M. 1987. Internal aeration and the

development of stelar anoxia in submerged roots: a multi-shelled

mathematical model combining axial diffusion of oxygen in the

cortex with radial losses to the stele, the wall layers, and the

rhizosphere. New Phytologist. 105: 221–245.

Armstrong, W.; Brandle, R.; Jackson, M.B. 1994. Mechanisms of

flood tolerance in plants. Acta Botanica Neerlandica. 43:

307–358.

Figure 4—Radial oxygen loss (ROL) in Spartina patens grown under

various soil reduction capacities while the reduction intensity was

maintained at –200 mV. Values followed by the same letter are not

significantly different at the 0.05 level. (Redrawn from Kludze and

DeLaune 1995.)

Figure 5—Carbon assimilation in Spartina patens grown under

various soil reduction capacities. The soil redox intensity was

maintained at -200 mV across all treatments. Data were collected at

17 days after growth in incubated soils. Values followed by the same

letter are not significantly different at the 0.05 level. OCE: Oxygen

Consumption Equivalent, mg O2 per kilogram soil. (Redrawn from

Kludze and DeLaune 1995.)
144

Armstrong, W.; Justin, S.H.F.W.; Beckett, P.M.; Lythe, S. 1991. Root

adaptation to soil waterlogging. Aquatic Botany. 39: 57–73.

Brix, H.; Sorrell, B.K. 1996. Oxygen stress in wetland plants:

comparison of de-oxygenated and reducing root environments.

Functional Ecology. 10: 521–526.

DeLaune, R.D.; Jugsujnda, A.; Reddy, K.R. 1999. Effect of root

oxygen stress on phosphorus uptake by cattail. Journal of Plant

Nutrition. 22: 459–466.

DeLaune, R.D.; Pezeshki; S.R. 1991. Role of soil chemistry in

vegetative ecology of wetlands. Trends in Soil Science. 1:

101–113.

DeLaune, R.D.; Pezeshki, S.R.; Lindau, C.W. 1998. Influence of soil

redox potential on nitrogen uptake and growth of wetland oak

seedlings. Journal of Plant Nutrition. 21: 757–768.

DeLaune, R.D.; Pezeshki, S.R.; Pardue, J.H. 1990. An oxidationreduction buffer for evaluating physiological response of plants to

root oxygen stress. Environmental & Experimental Botany. 30(2):

243–247.

DeLaune, R.D.; Smith, C.J.; Patrick, W.H. 1983. Relationship of

marsh elevation, redox potential and sulfide to Spartina

alterniflora productivity. Soil Science Society of America Journal.

47: 930–935.

Drew, M.C. 1990. Sensing soil oxygen. Plant Cell and Environment.

13: 681–693.

Drew, M.C. 1992. Soil aeration and plant root metabolism. Soil

Science. 154: 259–268.

Drew, M.C. 1997. Oxygen deficiency and root metabolism: injury

and acclimation under hypoxia and anoxia. Annual Review of

Plant Physiology and Plant Molecular Biology. 48: 223–250.

Drew, M.C.; Saglio, P.H.; Pradet, A. 1985. Larger adenylate energy

charge and ATP/ADP ratios in aerenchymatous roots of Zea

mays in anaerobic media as a consequence of improved internal

oxygen transport. Planta. 165: 51–58.

Faulkner, S.P.; Patrick, W.H.; Gambrell, R.P. 1989. Field techniques

for measuring wetland soil parameters. Soil Science Society of

America Journal. 53: 883–890.

Gambrell, R.P.; DeLaune, R.D.; Patrick, W.H. 1991. Redox processes in soils following oxygen depletion. In: Jackson, M.B.;

Davies, D.D.; Lambers, H., eds. Plant life under oxygen deprivation: ecology, physiology, and biochemistry. The Hague, The

Netherlands: SPB Academic Publishing: 101–117.

Gambrell, R.P.; Patrick, W.H. 1978. Chemical and microbiological

properties of anaerobic soils and sediments. In: Hook, D.D.;

Crawford, R.M.M., eds. Plant life in anaerobic environments. Ann

Arbor, MI: Ann Arbor Science: 375–423.

Hook, D.D.; Crawford, R.M.M. 1978. Plant Life in anaerobic

environments. Woburn, MA: Ann Arbor Science. 564 p.

Jackson, M.B.; Davies, D.D.; Lambers, H. 1991. Plant life under

oxygen deprivation: ecology, physiology, and biochemistry. The

Hague: SPB Publishing. 326 p.

Kludze, H.K. 1994. Gaseous exchange and wetland plant response

to soil redox conditions. Baton Rouge, LA: Louisiana State

University. 155 p. Ph.D. dissertation.

Kludze, H.K.; DeLaune, R.D. 1994. Methane emission and growth of

Spartina patens in response to soil redox intensity. Science

Society of America Journal. 58: 1838–1845.

Kludze, H.K.; DeLaune, R.D. 1995. Straw application effects on

methane and oxygen exchange and growth in rice. Science

Society of America Journal. 59: 824–830.

Kludze, H.K.; DeLaune, R.D.; Patrick, W.H. 1993. Aerenchyma

formation and methane and oxygen exchange in rice. Science

Society of America Journal. 51: 386–391.

Kludze, H.K.; Pezeshki, S.R.; DeLaune, R.D. 1994. Evaluation of

root oxygenation and growth in baldcypress in response to shortterm soil hypoxia. Canadian Journal of Forest Research. 24:

804–809.

Kozlowski, T.T. 1982. Water supply and tree growth. II: flooding.

Forestry Abstracts. 43: 145–161.

Kozlowski, T.T., ed. 1984a. Flooding and plant growth. New York:

Academic Press: 129–163.

Kozlowski, T.T. 1984b. Plant responses to flooding of soil. Bioscience. 34: 162–167.

Kozlowski, T.T. 1997. Responses of woody plants to flooding and

salinity. Tree Physiology Monographs. 1: 1–29.

Luxmoore, R.J.; Sojka, R.E.; Stolzy, L.H. 1972. Root porosity and

growth responses of wheat to aeration and light intensity. Soil

Science. 113: 354–357.

Mendelssohn, I.A.; McKee, K.L.; Patrick, W.H. 1981. Oxygen

deficiency in Spartina alterniflora roots: metabolic adaptation to

anoxia. Science. 214: 439–441.

Patrick, W.H.; DeLaune, R.D. 1972. Characterization of the oxidized

and reduced zones on flooded soil. Soil Science Society

American Louisiana. Estuarine Coastal & Marine Science. 4:

59–64.

Patrick, W.H.; DeLaune, R.D. 1977. Chemical and biological redox

systems affecting nutrient availability in the coastal wetlands.

Geoscience and Man: 131–137. Vol. 18.

Patrick, W.H.; Gambrell, R.P.; Faulkner, S.P. 1986. Redox measurements of soils. In: Methods of soil analysis. 3: chemical methods.

Book Ser. 5. Soil Science Society of America: 1255–1273.

Pezeshki, S.R. 1994. Plant responses to flooding. In: Wilkinson,

R.E., ed. Plant-environment interactions. New York: Marcel

Dekker, Inc.: 289–321.

Pezeshki, S.R.; Anderson, P.A. 1997. Responses of three bottomland woody species with different flood-tolerance capabilities to

various flooding regimes. Wetland Ecology and Management. 4:

245–256.

Pezeshki, S.R.; DeLaune, R.D. 1990. Influence of sediment

oxidation-reduction potential on root elongation in Spartina

patens. Acta Oecologia. 11: 377–383.

Pezeshki, S.R.; Matthews, S.W.; DeLaune, R.D. 1991. Root cortex

structure and metabolic response of Spartina patens to soil

redox conditions. Environmental & Experimental Botany. 31(1):

91–97.
145

Pezeshki, S.R.; Pardue, J.H.; DeLaune, R.D. 1993. The influence of

oxygen deficiency and redox potential on alcohol dehydrogenase

activity, root porosity, ehtylene production, and photosynthesis in

Spartina patens. Environmental & Experimental Botany. 33:

565–573.

Pezeshki, S.R.; Pardue, J.H.; DeLaune, R.D. 1996. Leaf gas

exchange and growth of flood-tolerant and flood-sensitive tree

species to soil oxygen deficiency. Tree Physiology. 16: 453–458.

Ponnamperuma, F.N. 1972. The chemistry of submerged soil.

Advances in Agronomy. 24: 29–96.

Ponnamperuma, F.N. 1984. Effects of flooding on soils. In:

Kozlowski, T.T., ed. Flooding and plant growth. Orlando, FL:

Academic Press Inc.: 1–44.

Reddy, K.R.; Ras, P.S.C.; Patrick, W.H. 1980. Factors influencing

oxygen consumption rates in flooded soil. Science Society of

America Journal. 44: 741–744.

Schat, H. 1984. A comparative ecophysiological study on the effects

of waterlogging and submergence on dune slack plants: growth,

survival, and mineral nutrition in sand culture experimens.

Oecologia. 62: 279–286.

Sorrell, B.K.; Brix, H.; Orr, P.T. 1993. Oxygen exchange by entire

root systems of Cyperus involucratus and Eleocharis

sphacelata. Journal of Aquatic Plant Management. 31: 24–28.

Teal, J.M.; Kanwisher, J.W. 1966. Gas transport in the marsh grass

Spartina alterniflora. Journal of Experimental Botany. 17:

355–361.

Van Wijck, C.; de Groot, C.J.; Grillas, P. 1992. The effect of

anaerobic sediment on the growth of Potamogeton pectinatus L.:

the role of organic matter, sulphide, and ferrous iron. Aquatic

Botany. 44: 31–49.

Vartapetian, B.B.; Jackson, M.B. 1997. Plant adaptations to

anaerobic stress. Annals of Botany. 79 (Supp. A): 3–20.

Williams, W.T.; Barber, D.A. 1961. The functional significance of

aerenchyma in plants. S. Exper. Bot. Sym. 15: 132–144.
146

RIVERBANK RESTORATION IN THE SOUTHERN UNITED STATES:

THE EFFECTS OF SOIL TEXTURE AND MOISTURE REGIME ON

SURVIVAL AND GROWTH OF WILLOW POSTS

S. Reza Pezeshki, Steven D. Schaff, and F. Douglas Shields, Jr.1

Abstract—Field studies were conducted to quantify the relationship between soil conditions and growth of black willow

posts planted for riverbank erosion control along Harland Creek (HC) and Twentymile Creek (TC) sites in Mississippi. Both

sites had a wide range of soil texture and moisture regimes. Soil texture, water level, redox potential (Eh), and willow

survival and growth were monitored. At the HC site, growth was lower in posts located at elevational extremes than for

posts planted approximately 0.5 m above creek elevation at baseflow. Optimum conditions for growth were provided at

moderate elevations characterized by groundwater levels that fluctuated around 50 cm beneath the soil surface. Data from

the TC site also indicated a close correlation between soil texture, moisture, and survival of willows. Posts grown in siltyclay soils displayed low survival and growth in comparison to those grown in sandy soils. Locations along the creek

characterized by sandy soils, adequate soil moisture, and well-drained conditions resulted in high survival and growth.

INTRODUCTION

Riverbanks and stream reaches in the Southern United

States are often subject to accelerated erosion and

associated habitat deterioration. However, there are

opportunities for both erosion control and improvement of

natural habitat. Many techniques are presently used for

riverbank restoration including planting vegetation primarily

composed of woody species on eroding banks. Willow posts

(Salix spp.) have been extensively utilized for such

restoration (Roseboom 1993; Shields and others 1995a,

1995b; Watson and others 1997). Large-diameter (> 7.5 cm)

willow cuttings with lengths up to 4 m are sometimes used

to stabilize rapidly eroding banks. The deeply planted posts

provide mechanical control of erosion during the period of

initial growth and establishment. A dense array of growing

willow posts is intended to modify riparian habitats, fostering

natural succession to a diverse riparian plant community

and relatively stable banks. Despite reported success on

many restoration sites, low survival rates have been

reported in certain areas. For instance, survival rates of > 40

percent by the end of the first growing season have been

reported for northern Mississippi (Shields and others

1995b). This has been attributed to many factors including

post location on the bank, parasites, flooding, drought, and

soil texture (Shields and others 1998). Field observations

suggest that regimes of excess and deficit soil moisture may

be primary factors. Experimental evidence on adverse

extreme soil moisture effects have come primarily from a

greenhouse study conducted on black willow (Salix nigra

Marsh.) cuttings and have shown flooding or drought have

significant adverse effects on physiological functioning,

growth, and biomass production (Pezeshki and others

1998).

Field observations also suggest the potential effects of soil

moisture on the growth of willow cuttings. Willow posts

planted on steep banks appeared to do better in middle

elevation as compared to those planted close to the river or

at higher elevations further from the river (Shields and

others 1998). At the two elevational extremes, willows

experience quite different situations. At high elevations,

periodic soil drought may impose severe water stress to

which the posts are likely to be sensitive. In contrast, at low

elevations close to the permanent water table, willows are

likely to experience frequent root inundation due to soil

flooding. In such conditions, willows will be subjected to soil

and, thus, root oxygen deficiency as well as to soil chemical

changes that result from flooding (Pezeshki 1994). The

effects of soil oxidation-reduction conditions (redox potential,

Eh) could be significant because reducing soil conditions

(low soil Eh) may impose substantial stress on plants

including willow, a species considered to be flood tolerant

(Hook 1984). Depending on the level of soil Eh in the root

zone, significant reductions in root growth can occur in many

flood-tolerant woody species (Pezeshki 1991). In addition,

substantial alterations in normal root metabolic activities

may occur leading to the disruption of various root

processes critical to a plant’s functioning. Such processes

include water and nutrient uptake, production of metabolites,

and oxidation of the immediate rhizosphere under flooded

conditions (Pezeshki 1994). Partial or complete death of

roots may occur leading to insufficient masses of functional

roots, and this may create root/shoot imbalances. Such

imbalances may initiate a chain of events leading to shootwater deficits and massive root and shoot death (Kozlowski

1982, 1997; Pezeshki 1991, 1994).

Soil texture is also critical because soil-pore space,

moisture-holding capacity, and oxidation-reduction capacity

are closely correlated with texture. Poor performance of

willow posts planted on streambanks composed of fine soils

has been observed in the field (Abt and others 1996). Posts

1

 Pezeshki, Professor of Biology, and Schaff, Graduate Student, Department of Biology, The University of Memphis, Memphis, TN 38152;

Shields, Research Hydraulic Engineer, National Sedimentation Laboratory, USDA-ARS, Oxford, MS 38655.

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability of

wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.
147

also could experience severe drought in coarse soils (sandgravel) on steep banks during the low-precipitation periods

due to low moisture-holding capacity of coarse soils. Thus,

both soil-moisture regime and texture are inter-related and

may play a critical role in the success of willow posts planted

on a restoration site.

Clearly, the poor performance of willow posts involves a

complex set of environmental factors to which the posts are

responding. The environmental conditions and the

associated plant responses change substantially over the

course of a typical growing season in the highly dynamic

riverbank systems. Therefore, the present studies were

designed to quantify the relationship between soil texture

and moisture regime and physiological functions, survival,

and growth of willow posts under field conditions.

MATERIALS AND METHODS

Field work was conducted on black willow (Salix nigra

Marsh.) posts planted for streambank restoration at two

restoration sites in northern Mississippi, the Harland Creek

(HC) restoration site, and the Twentymile Creek (TC)

restoration site (fig. 1).

Harland Creek Restoration Site

The HC restoration site was located in Holmes County, MS

(fig.1). Black willow posts were harvested from local

populations during the dormant season in 1994. The posts

were approximately 3 m in length with a minimum diameter

of 7.5 cm at the base. Posts were placed in holes 2.4 m

deep, thus, approximately 0.6 m of the post remained

aboveground. Willow posts were spaced at 0.9 m-centers for

four rows of posts. The first row was placed 0.6 m from

where the stream channel met the bank at the time of

planting. Posts were planted along the bank in February

1994. In 1996, a total of 13 posts were selected along an

elevational gradient extending from the upstream and ending

at the last downstream post. These posts were located at

various distances from the creek. PVC piezometers with a

diameter of 3.1 cm were established in proximity of each

post. Measurement locations were also established at each

piezometer to monitor soil Eh. Water, soil, and plant

measurements (described in Methodologies) were continued

through the end of the 1996 growing season (November).

Twentymile Creek Restoration Site

The TC restoration site, located in Lee County, MS (fig. 1),

was the subject of intensive data collection. Black willow

posts were planted according to the methodology described

above in February 1998. At this site, four transects were

established along the creek. Each transect was placed

perpendicular to the creek and extended from the toe of the

creek to the edge of the planting zone, approximately half

way up the bank. Three plots were located within each

transect: one at each end of the elevational extremes and

one in the middle elevation. Piezometers were installed in

the middle of each plot to monitor the water level below the

soil surface. Each plot consisted of 12 posts. The posts were

grouped into 2 rows each consisting of 6 plants, parallel to

the stream for a total of 12 posts per plot. Soil and plant

measurements were initiated in the spring of 1998 and

continued through the end of the growing season.

METHODOLOGIES

The frequency and duration of flooding of both sites was

analyzed using stage data collected at 15-minute intervals

from nearby U.S. Geological Survey gauging stations. The

TC transects were about 0.4 to 1.7 km downstream from the

gauge, whereas the HC site was about 6.3 km upstream

from the gauge. The applicability of the HC gauge data to

our site was verified by comparing the signal from the U.S.

Geological Survey gauge with a temporary gauge located

about 1 km upstream from our site. Stages were transformed

into height above baseflow elevation for assessing impacts

at each site.

At both sites, soil samples were collected from 15, 30, 60,

and 90 cm below the soil surface during piezometer

installation. These samples were used to characterize the

soil texture for each post location using standard methods

(Brady 1974). For the HC site, the percentage of sand, silt,

and clay are approximate as determined by the “feel

method.” For the TC site, soil particle size distribution was

determined by the “feel method” and by the USDA

Agricultural Research Service, Midsouth Area, using a laser

scattering particle size distribution analyzer (Horiba, Model

LA-910).

At each site, soil redox potential (Eh) was measured at a

distance of 15 cm from each well at 15-, 30-, 60-, and 90-cm

depths below the soil surface using platinum-tipped redox

electrodes, a millivoltmeter (Orion, Model 250A), and a

calomel reference electrode (Corning, Model 476350).

Measurements were recorded after the redox probes had

been in the ground at least for 2 hours to allow for

equilibration of the electrodes. Measurements of Eh were

replicated (two measurements at each depth) per sampling

location per sampling date. The water level at each

monitoring well was quantified using a measuring tape.

Twentymile Creek

restoration site

Harland Creek

restoration site

Figure 1—Map of the United States showing locations of the

Harland Creek and Twentymile Creek restoration sites in northern

Mississippi, USA.
148

Plant measurements at each site included selected

physiological measurements and growth responses. Gas

exchange measurements (leaf conductance and net carbon

fixation) were conducted approximately twice each month

using a portable photosynthesis system (CIRAS1, PP

Systems, England). Measurements were conducted on

mature, intact, attached leaves located on the upper onethird of the branches on each post between the hours of

11:00 and 14:00 during each site visit. The environmental

factors of air temperature, leaf temperature, and

photosynthetic photon flux density (PPFD) were recorded

followed by measurements of net photosynthesis (Pn) and

stomatal conductance (gw) on each sample leaf. Replicated

sample leaves were used for the measurements during each

field visit.

The growth of posts at the HC site was quantified by

biomass sampling at the end of the growing season. Posts

were destructively sampled and dry weight of biomass

components was determined (Shields and others 1998).

Growth of posts at the TC site was determined by

measurement of height on study posts at the end of the

1998 growing season. No destructive samples were

conducted on these posts because the posts were to be

utilized in a longer study. The general linear models (GLM)

procedure of the Statistical Analysis System was used to

test for differences in means among different posts planted

at various elevations relative to the creek. Regressions for

the relationships between post survival, height growth, and

soil texture were calculated using the SAS System

(Statistical Analysis Systems 1994).

RESULTS

Harland Creek Restoration Site

Posts were planted at the HC site in February 1994, and we

conducted measurements during the 1996 growing season.

Rainfall and runoff patterns during the period 1994–96 were

within limits typical of the period of record. However, in 1994

precipitation and streamflow were well above average during

late summer. During our study, inundation of the posts was

brief and infrequent. The median duration of flooding ranged

from about 26 hours for the lowest posts to 15 hours for the

highest posts. Median times between flooding ranged from

more than 8 to more than 13 days, depending on elevation.

At the HC site, study posts were grouped into four distinct

groups based on soil characteristics, relative elevation from

the creek, and the distance to ground water (table 1). Posts

in the upper part of the bank (group I) where soil texture was

sandy gravel experienced little or no soil flooding as was

evidenced by soil Eh data (> 350 mV). However, due to

location and soil texture, the posts were likely to have

encountered drought. Net carbon fixation indicated a

significantly lower mean as compared to the posts that were

located in the middle slopes (group II, percent III). No

biomass sampling was done for this group, but gas

exchange data suggest that these posts were stressed.

Groups II and III consisted of posts located in the middlebank elevation as verified by elevation and depth to water

table (table 1). These posts had the highest measured net

carbon assimilation rates among the study groups. The two

groups showed no significant differences in gas exchange

responses, but the mean (root and shoot) biomass per post

was significantly greater in group II than group III (table 1).

The difference in the observed responses between the two

groups was attributed to the difference in soil texture that

was more favorable for group II (less cohesive, higher Eh)

than group III (more cohesive, lower Eh). Group IV posts

were located at low elevations in sand-gravel soil. The water

table was higher for this group as compared for the middlebank groups; thus, soil Eh was in the reducing range.

However, it was not significantly greater in intensity

compared to the middle-bank groups. This finding was

primarily attributed to the sand-gravel soil texture, which had

little capacity for intense reduction. Examination of root

Table 1—Field data of post location, soil texture, depth to water table, soil Eh,

net carbon fixation, root dry biomass, and shoot dry biomass for posts studied

at Harland Creek restoration site, Mississippi

 Post location on the streambank

Lower bank Medium Medium Upper bank

Variable group IV group III group II group I

Soil texture Sand-gravel Sand-silt-clay Sand-silt Sand-gravel

Water depth (m) 0.33 c 0.58 b 0.54 b 1.13 a

Soil Eh (mV) 202 b 107 b 172 b 384 a

Net carbon fixation

(μmol CO2

 m-2 leaf s-1) 8 b 10 a 11 a 7 b

Root dry mass (g) 100 29 b 192 a NA

Shoot dry mass (g) 24 173 b 1029 a NA

NA = missing data.

Different letters within each column indicate statistically significant differences

(p< 0.05).
149

distribution patterns showed that for posts located at low

elevations characterized by low soil Eh conditions, 100

percent of roots excavated were located between a depth of

0 to 15 cm. Posts at medium elevation showed substantially

different root-growth patterns depending on the soil texture.

Group II located on sand-silt soil had significantly higher root

biomass than group III located on soils that contained some

clay (fig. 2). In addition, roots were more uniformly

distributed at various depths in soil profile in group II

whereas roots were confined to the top 15 cm of the soil in

group III. Overall, conditions for growth of willow posts at HC

site were best at moderate elevations characterized by

sand-silt soil texture and water levels that fluctuated close to

the soil surface allowing for adequate soil moisture but

frequent drainage. In contrast, posts located at higher

elevations relative to the baseflow stage suffered from soil

moisture deficits whereas posts located at the bank toe were

hampered by flooded soils and low soil Eh conditions.

Twentymile Creek Restoration Site

At the TC site, the study plots on each transect were located

along an elevational gradient. At each transect, the lowelevation plot was flooded frequently due to proximity to the

creek, which rose whenever light to moderate rains fell. The

middle plot flooded at a moderate frequency but required

moderate to heavy rainfall throughout the watershed

whereas the high-elevation plot was located at the top of the

planting zone and was never flooded. Rainfall and runoff

patterns during the 1998 growing season were typical of the

period of record. Plots located 0.3 m above baseflow were

flooded about 10 percent of the time during March through

July (later data not available at this writing), whereas those

at intermediate elevations (+ 1.0 to 1.5 m above base stage)

were flooded only about 1 percent of the time, and those at

highest elevations (>3 m above base stage) were not

flooded during this period. Flooding episodes lasted only 2

to 39 hours, and median times between flooding ranged

from 8 to 26 days, depending on elevation.

The soils in each of the elevational categories were fairly

similar in texture ranging between 76 to 62 percent coarse

sandy materials (table 2). The low-elevation plots were

closest to the water table (0.3 m) and had moderately

reduced soils due to the saturated conditions (Eh +230 mV).

In contrast, the medium- and high-elevation plots had

aerated soils with Eh values of +480 and +520 mV,

respectively (table 2).

Height growth, survival rates, and gas-exchange

measurements showed no significant differences across the

elevational groups. However, posts in the medium-elevation

plots grew 98.4 cm on average, 196 percent more than the

50.2-cm average for low-elevation plots and 154 percent

more than the 63.8 cm recorded for the high-elevation plots.

The high-elevation plots had the highest survival rate at 67

percent followed by the middle plots at 50 percent and the

low plots at 46 percent (table 2). However, differences in

growth and survival were not statistically significant.

The soil texture effects were apparent when plots of the

same elevation were compared between reach 1 and reach

3. The amount of fine-grained soil (silt and clay) in reach 3

plots was double that of reach 1 plots at all elevations. The

depth to the water table showed no significant difference

across the lowest plots; however, there were slight

differences (0.5 and 0.3 m, respectively) between the

middle- and high-elevation plots (table 3). This difference

may be biologically insignificant, however, because this is

well below the root establishment zone. Similarly, the only

Figure 2—Total root biomass distribution for willow posts in soil

profile at three depths: 0 to 15 cm, 15 to 30 cm, 30 to 60 cm, for

posts harvested at Harland Creek site, Mississippi. Data for group II

represent the mean for two posts while data of group III represent

the mean for four posts.

Table 2—Field data of post location, soil texture, depth to

water table, soil Eh, height growth, survival, and net

carbon fixation for willow posts studied at the Twentymile

Creek restoration site, Mississippi

Post location on the streambank

Lower Upper

Variable bank Medium bank

Soil texture

(percent coarse) 70 62 76

Water depth (m) 0.3 a 1.4 b 1.7 c

Soil Eh (mV) 230 a 480 b 520 c

Height growth (cm) 50.2 a 98.4 a 63.8 a

Survival (percent) 46 a 50 a 67 a

Net carbon fixation

(μmol CO2

 m-2 leaf s-1) 11.7 a 12.3 a 10.3 a

Means in each row followed by the same letter are not significantly

different across groups at the p< 0.05 level.
150

plots to have biologically significant differences in soil Eh

were the low plots. The low plot in reach 1 had a significantly

lower Eh value than in reach 3, even though there was only

a 3-cm difference in the depth to the water table, and the

majority of the soil was sand, which characteristically has a

lower capacity for reduction. Nevertheless, both plots were in

the anaerobic range (below + 350 mV). All of the middle and

high plots had Eh values indicative of aerated soils (above +

350 mV) (table 3).

Gas exchange data showed a consistent pattern of

response to soil texture. Net photosynthetic rates were

consistently higher in sandy soils; however, rates were

significantly higher only in high-elevation plots. In these

plots, average carbon fixation was 40 percent higher in

coarse-grained sandy soils than in fine soils (table 3). The

middle plots showed moderately significant results (p =

0.0545) with an average increase in carbon fixation of over

30 percent. Average height growth per post followed similar

trends (fig. 3). The low-elevation plots failed to show a

significant difference in growth; however, posts grown in

sandy soils were considerably taller than posts grown in silt

and clay soils (139 cm and 77 cm, respectively). Posts

grown in coarse soils showed a significantly greater amount

of growth at both the middle- and high-elevation plots. Posts

grown in sandy soils in the middle-elevation plots showed a

threefold increase in growth (246 cm compared to 74 cm)

and a fourfold increase at the high-elevation plots (131 cm

compared to 31 cm) (fig. 3). Post survival showed the most

consistent significant effects of soil texture as posts grown in

sandy soils had greater survival in all plots. Posts planted in

predominantly sandy soils had mean survival rates of 67

percent, 75 percent, and 92 percent, for low-, middle-, and

high-elevation plots, respectively. In contrast, posts in finegrained soil (silty/clay soils) had significantly lower survival

rates of 25 percent, 25 percent, and 42 percent, for the

same respective plot elevations (fig. 3). The relationship

between willow survival and soil conditions indicated that

Table 3—Field data comparing the effects of soil texture on net carbon fixation

when plot elevation (thus, depth to water table and soil redox potential) is held

constant. Data represents mean values taken over the first growing season for

depth to water table, soil Eh, and net carbon fixation for willow posts studied at

Twentymile Creek restoration site, Mississippi

Silt + clay Sand H2O depth Soil Eh Pn

- - - - Percent - - - m mV μmol CO2 m-2 leaf s-1

Low plots Reach 1 23 77 0.26 a 146 a 12.4 a

Reach 3 48 52 .31 a 259 b 11.0 a

Middle plots Reach 1 29 71 1.2 a 378 a 13.9 a

Reach 3 67 33 1.7 b 376 a 9.7 a

High plots Reach 1 11 89 1.8 a 553 a 11.7 a

Reach 3 30 70 1.5 b 491 b 7.0 b

Means in each column for each elevation followed by the same letter are not significantly different

across groups at the p< 0.05 level.

Figure 3—Survival and height growth of willow posts planted on

fine vs. coarse soils at the Twentymile Creek site, Mississippi.

Data represents the mean height growth per post (cm) and

percent survival per plot for posts grown at similar elevation

with varying soil textures. Sandy soils were consistently

observed at reach 1 in comparison to the silt and clay soils

observed at reach 3 (for specific proportions of each soil

texture see table 3). Means for each elevation with similar

letters represent no significant difference at the p <0.05 level.
151

posts appear to have low-survival rates in soils containing a

high percentage of fine particles (silt plus clay) within the

upper soil layer of 0 to 30 cm in depth. Height growth was

also sensitive to a high percentage of fine particles.

DISCUSSION

The results from both restoration sites demonstrated the

complexity of soil-plant interactions in highly dynamic creek

banks such as the HC and the TC restoration sites that are

characterized by steep slopes, a wide range of soil moisture

conditions, and various soil textures. Willow posts displayed

a great level of sensitivity to drought and flooding (low soil

Eh conditions). For instance, on the HC site, soil Eh for posts

in group I was above + 350 mV, thus, in the aerated range,

suggesting that these posts were not exposed to flooding or

soil saturation during any of the measurement days (table 1).

Nevertheless, this group had the lowest gas exchange rates

relative to other groups. This finding was attributed to

periodic drought encountered by this group due to elevation

and the poor water-holding capacity of the soil (sand-gravel).

Similar results were obtained in a greenhouse study where

posts under drought treatment had low photosynthetic rates

(Pezeshki and others 1998). In contrast, posts in group IV

were at the lowest elevations to the creek relative to other

groups. Here, water levels were high (close to the soil

surface) creating reducing soil conditions indicated by low

Eh values (+202 mV). This group had significantly lower gas

exchange rates relative to other groups (II and III). In a

greenhouse study, willow posts subjected to a continuously

flooded condition had significantly lower net photosynthesis

compared to control (Pezeshki and others 1998). Reducing

soil Eh can adversely affect photosynthetic rates in species

considered to be flood tolerant (Pezeshki 1994). Field root

biomass data indicated that reductions in soil Eh had a

negative impact on root biomass. Root development is

slowed in areas of low soil Eh conditions because of the lack

of oxygen, which is needed to carry out normal root

respiration. In a greenhouse study, little or no root growth

was found for black willow posts at locations that remained

subjected to continuous flooding and soil Eh around - 200

mV (Pezeshki and others 1998). This lack of roots indicated

that willow posts either failed to initiate roots or that such

roots, if initiated, died in response to the continued, intense

soil-reducing conditions. Other studies have shown that root

biomass in black willow was significantly decreased as

flooding depth and duration increased (Donovan and others

1988).

Black willow is a flood-tolerant species; thus, it possesses

physiological/morphological adaptations allowing oxygen

transport to the root zone. However, as soil Eh decreases,

the chemical and biological demand for oxygen within the

soil rises. Thus, the oxygen-delivery system may be

overwhelmed by such increasing demand that is

concomitant with increased internal demand for oxygen

(Pezeshki 1994). The lower gas exchange rates found in

posts under drought and flooding on HC site partially explain

the observations on changes in vigor and survival of willow

posts due to elevation reported by Abt and others (1996).

These observations agreed that posts planted along an

elevational gradient from the creek differed in survival rates.

Posts located close to the creek at similar elevations to the

creek and exposed to prolonged, frequent flooding had

mortality rates of 80 percent. The posts at the highest

elevations and greatest distance from the creek, which were

likely to experience droughty conditions, had mortality rates

of about 91 percent. However, higher survival rates of 39 to

58 percent were associated with the posts located in the

area between these zones (Abt and others 1996).

The TC data further indicated the significant influence of soil

texture on willow performance. Net carbon fixation

substantially decreased for posts grown at middle to high

elevations in coarse-grained sandy soils (table 3). Soil

texture also had a substantial impact on height growth and

survival as they were both significantly greater in coarse

sandy soils than in fine silt and clay soils (fig. 3). Regression

analysis indicated a significant negative correlation (r = -

0.68 p = 0.0438) between the percentage of fine particles

(silt + clay at 15- and 30-cm depths) and survival. Height

growth was also significantly correlated (r = - 0.69 p =

0.0412) to the average percentage of fine particles (at 15-

and 30-cm depths) showing low growth at soils that

contained a high percentage of fine particles (silt + clay).

CONCLUSIONS

At both restoration sites, soil texture and moisture regime

were important factors affecting survival, physiological

functions, and growth of willow posts. For instance, much of

the variations in survival and growth were due to soil texture

represented by the percentage of fine particles (silt + clay) in

the upper 30 cm of the soil. High survival and growth in

willow posts required ample soil moisture (but

nonwaterlogging conditions) and adequate drainage in the

upper soil layer, approximately the top 45 to 60 cm of the

soil. In addition, under favorable soil-moisture conditions,

posts perform better in soils characterized by coarsetextured rather than fine-grained soils. From a restoration

perspective, use of willow posts for riverbank restoration

projects is intended to help stabilize the banks while

improving habitat quality. Thus, the use of willow posts

remains a low-cost restoration strategy; however, a proposed

planting site must be evaluated prior to the undertaking of

large-scale restoration. Many factors must be considered

including soil texture, moisture regime, slope, distance to the

creek, and soil oxidation-reduction conditions in order to

improve the prospect for high survival and growth of willow

posts.

ACKNOWLEDGMENTS

This work was supported in part by a grant from the U.S.

Department of Agriculture, National Sedimentation

Laboratory, Cooperative Agreement #58–6408–6–029.

Graduate students Christopher Brown, Mitch Elcan, and

Jason Farmer helped with field data collection. Sara Davis

and Michael Eubanks read an earlier version of this paper

and made many helpful comments.
152

REFERENCES

Abt, S.R.; Watson, C.C.; Burgi, J.P. [and others]. 1996. Willow posts

construction and evaluation on Harland Creek, Mississippi. In:

Proceedings of the sixth Federal interagency sedimentation

conference: II14–II23.

Brady, N.C. 1974. The nature and properties of soils. New York:

McMillan Publication Co. 639 p.

Donavan, L.A.; McLeod, K.W.; Sherrod, K.C.; Stumpff, N.J. 1988.

Response of woody swamp seedlings to flooding and increased

water temperatures. I: growth, biomass, and survivorship.

American Journal of Botany. 75: 1181–1190.

Hook, D.D. 1984. Waterlogging tolerance of lowland tree species of

the South. Southern Journal of Applied Forestry. 8: 136–149.

Kozlowski, T.T. 1982. Water supply and tree growth. II: flooding.

Forestry Abstracts. 43: 145–161.

Kozlowski, T.T. 1997. Responses of woody plants to flooding and

salinity. Tree Physiology Monographs. 1: 1–29.

Pezeshki, S.R. 1991. Root responses of flood-tolerant and floodsensitive tree species to soil redox conditions. Trees. 5: 180–186.

Pezeshki, S.R. 1994. Plant responses to flooding. In: Wilkinson,

R.E., ed. Plant environment interactions. New York: Marcel

Dekker, Inc.: 289–321.

Pezeshki, S.R.; Anderson, P.H.; Shields, F.D. 1998. Effects of soil

moisture regimes on growth and survival of black willow (Salix

nigra) posts (cuttings). Wetlands. 18: 460–470.

Pezeshki, S.R.; Shields; F.D., Jr.; Anderson, P.H. 1998. The

relationships between soil conditions and growth of willow posts

on streambanks. In: Abt, S.R.; Young-Pezeshk, J.; Watson, C.C.,

eds. Proceedings of the international conference on water

resources engineering. New York: American Society of Civil

Engineers: 447–452.

Roseboom, D. 1993. Breakwater installation and vegetative

stabilization in Illinois. In: Proceedings, U.S. Army Corps of

Engineers workshop on reservoir shoreline erosion. Publ.

W–93–1. Vicksburg, MS: Army Corps of Engineers, Waterways

Experiment Station: 166–173.

Statistical Analysis Systems. 1994. Procedures guide. Version. 6. 3rd

ed. Cary, NC: SAS Institute Inc. 705 p.

Shields, F.D., Jr.; Cooper, C.M.; Knight, S.S. 1995a. Experiments in

stream restoration. Journal of Hydraulic Engineering. 121:

494–502.

Shields, F.D., Jr.; Knight, S.S.; Cooper, C.M. 1995b. Streambank

protection and habitat restoration. In: Espey, W.H., Jr.; Combs,

P.G., eds. Water resources engineering; Proceedings of the first

international conference on water resources engineering. New

York: American Society of Civil Engineers: 721–725.

Watson, C.C.; Abt, S.R.; Derrick, D. 1997. Willow post for bank

stabilization. Journal of American Water Research Association.

33: 293–3.
153

A firm understanding of the driving forces controlling

variation among wetland forests continues to elude

scientists and land managers—specifically the

biogeochemical processes controlling vegetation production.

Within contrasting wetland forests, insight into the

biogeochemical processes driving productivity levels may be

found by examining the degree to which nitrogen and

phosphorus are balanced within the wetland vegetation.

Lockaby and Conner (1999) suggest that there exists a

biogeochemical continuum for wetland forests based on the

relationship between N:P ratios. Koerselman and Meuleman

(1996) have indicated that N:P ratios in forest vegetation

may serve as a measure of the biogeochemical constraints

on vegetation net primary production (NPP). Thus, the

position of a particular wetland forest on this N:P continuum

reflects the integration of its geomorphic position and

biogeochemical history and may have a predictive value with

regard to levels of NPP (Lockay and Conner 1999).

Ultimately, the synchrony of nutrient availability and plant

uptake influences the levels of NPP within wetland forests.

In the Southeastern United States, riverine systems are

often broadly characterized as either red-water or blackwater systems. Red-water rivers have their origins in the

Piedmont and are characterized hydrochemically as rich in

nutrients and suspended organic matter. Conversely, blackwater rivers, having their origins within the Coastal Plain, are

poorer in both. Within these systems, the contrasting quality

of detrital inputs influences the mineralization and

immobilization patterns governing nutrient availability.

The purpose of this study is to examine and characterize the

biogeochemical processes influencing nutrient availability

and levels of NPP within two wetland forest types. The river

systems being used for this study are the Altamaha River, a

red-water river system, and the Satilla River, a black-water

river system. In 1999, annual litterfall mass for the Altamaha

and Satilla Rivers totaled 8.3 and 6.1 t per hectare per year,

respectively. Other data to be presented from this study

include: aboveground woody biomass, litter decomposition,

microbial biomass C, N, and P, retranslocation efficiency,

and soil nutrient availability.

Data to be presented for each of the aforementioned foci are

the year-one results of an ongoing study.

REFERENCES

Koerselman, W.; Meuleman, A.F.M. 1996. The vegetation N:P ratio:

a new tool to detect the nature of nutrient limitation. Journal of

Applied Ecology. 33: 1441–1450.

Lockaby, B.G.; Conner, W.H. 1999. N:P balance in wetland forests:

productivity across a biogeochemical continuum. The Botanical

Review. 65(2): 171–185.

DIFFERENCES IN NET PRIMARY PRODUCTION AND

BIOGEOCHEMISTRY BETWEEN CONTRASTING FLOODPLAIN FORESTS

Erik B. Schilling and B. Graeme Lockaby1

1 Schilling, Research Associate, and Lockaby, Professor, Auburn University, School of Forestry and Wildlife Sciences, Auburn, AL 36849-5418.

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability of

wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.
154

INTRODUCTION

Agriculture is a source of sediment, nutrient, and pesticide

input into surface waters (Cooper 1993). Rainfall-related

runoff and spraydrift are two major sources for the

agricultural nonpoint-source pollution of surface waters

(Groenendijk and others 1994). The fact that nonpointsource pollutants are recognized as the single greatest

threat to surface waters (Loague and others 1998) has

heightened concerns about sustainable agriculture and has

highlighted the need for measures to minimize input and risk

(Mainstone and Schofield 1996).

Constructed wetlands have been suggested and used as a

potential risk-reduction strategy for nonpoint-source pollution

(Hammer 1992, Mitsch 1992, Van der Valk and Jolly 1993).

Whereas the fate and retention of nutrients and sediments in

wetlands is understood quite well (Hammer 1992, Van der

Valk and Jolly 1993), the same cannot be said of

agrochemicals (Baker 1993). Only very few studies refer to

the potential of wetlands for removal of herbicides and some

other organic chemicals (Kadlec and Hey 1994, Lewis and

others 1999, Wolverton 1987). Since wetlands have a high

ability to retain and process material, it seems reasonable

that constructed wetlands as buffer strips between

agricultural areas and receiving surface waters could

mitigate the impact of pesticides in this runoff (Rodgers and

Dunn 1993).

During the last decades, a shift to lower water quality in

Western Cape rivers has been observed. This shift has also

occurred in the middle and lower reaches of the Lourens

River and is attributed to intensified agriculture, sediment

input, and loss of indigenous vegetation (Tharme and others

1997). No information is available on the extent to which

toxic substances are responsible for the degradation of the

Lourens River. In order to minimize the input of sediment into

the Lourens River, a small flow-through constructed wetland

was built in 1991 into one of the tributaries. Constructed

wetlands have been used so far in South Africa mainly for

wastewater treatment (Wood 1990).

The aim of the present study is to assess the effectiveness

of a small constructed wetland for control of nonpoint-source

pesticide input into the Lourens River. Retention is assessed

for particle-associated pesticide input via runoff and for

water-diluted pesticide input following drift during spray

application. Sediment- and nutrient-trapping efficiencies are

evaluated, and the effects of the wetland in terms of toxicity

alteration are addressed using an in situ exposure bioassay.

MATERIALS AND METHODS

Study Region and Climate Conditions

The Lourens River rises at an altitude of 1080 m in the

Hottentots Holland Nature Reserve and flows in a

southwesterly direction for about 20 km before discharging

into False Bay at Strand (34°06 ́ S., 18°48 ́ E.). The river is

surrounded by intensive farming activities (orchards and

vineyards) in its middle reaches after leaving a naturally

vegetated fynbos area and before flowing through the town

of Somerset West. The Lourens River has a total catchment

area of approximately 92 km2 and receives a mean annual

rainfall of 915 mm.

The orchards mainly consist of pears, plums, and apples

(total growing area: 4 km2). The pesticide application period

starts in the study area ́s orchards in early August and ends

at the end of January (pears) or end of March (apples).

Organophosphorous (OP) insecticides like azinphos-methyl

and chlorpyrifos are used between October and February

quite frequently on pears and plums. Endosulfan is applied

mainly to apple orchards.

USE OF A CONSTRUCTED WETLAND TO REDUCE NONPOINT-SOURCE

PESTICIDE CONTAMINATION OF THE LOURENS RIVER, SOUTH AFRICA

Ralf Schulz1

1 Supervisory Ecologist, Department of Zoology, Private Bag X1, University of Stellenbosch, Matieland 7602, South Africa. Present address:

Zoological Institute, Technical University, Fasanenstrasse 3, D-38092 Braunschweig, Germany, R.Schulz@tu-bs.de

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability of

wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.

Abstract—The Lourens River, Western Cape, South Africa, and its tributaries situated in an intensively cultivated orchard

area receive pesticide contamination during rainfall-induced runoff and during spraydrift. A 0.44-ha constructed wetland,

built in 1991 in one of the tributaries (summer flow 0.03 m3 per second), was studied in order to assess its effectiveness in

reducing nonpoint-source agricultural pesticide contamination. Even high levels of particle-associated azinphos-methyl (43

μg per kilogram), chlorpyrifos (31 μg per kilogram) and prothiofos (6 μg per kilogram) introduced via runoff were not

detectable in the outlet suspended-particle samples. Recovery of water-diluted azinphos-methyl contamination at inlet

concentrations of about 0.55 μg per liter during spraydrift indicated much less retention, approximately 51 percent. Comparison of in situ bioassays of bloodworms (Chironomus spec.) above and below the wetland revealed a reduction of contamination in terms of toxicity from 41 to 2.5 percent. TSS, ortho-phosphate, and nitrate were retained in the wetland with

trapping efficiencies of 78, 75, and 84 percent, respectively.
155

The constructed wetland studied in the present investigation

is located in one of the tributaries shortly before its entry into

the Lourens River. This tributary has an average width and

depth of 0.89 m by 0.30 m and a current velocity of about

0.1 m per second. Average discharge in the tributary is 0.03

m3 per second in January and 0.32 m3

 per second in July.

The study period lasted from December 1998 until June

1999. A heavy rainfall event resulting in edge-of-field runoff

during the pesticide application period occured on

December 14, 1998. The period between January and the

middle of April was characterized by low rainfall (dry period),

whereas the following months were again coupled with high

rainfall (wet period).

Separate studies revealed that the Lourens River receives

contamination from the surrounding farming areas via

rainfall-related runoff and via spraydrift during spray

application. Runoff produced concentrations of particleassociated chemicals as high as 244 μg per kilogram

azinphos-methyl, 69 μg per kilogram chlorpyrifos and 245 μg

per kilogram total endosulfan. Spraydrift-related pesticide

input elevated the levels of water-diluted pesticide in the

Lourens River to about 0.04 μg per liter azinphos-methyl,

with peak concentrations at about 1.7 μg per liter in those

tributaries receiving spraydrift directly. Effectiveness of

pesticide retention in the constructed wetland was studied

for both scenarios—runoff and spraydrift.

Description of the Constructed Wetland

The constructed wetland was built in 1991 into one of the

southeastern tributaries draining the surrounding farmland to

prevent nonpoint-source pollution with suspended particles

from entering the Lourens River. The catchment area of the

wetland comprises about 15 ha of orchard, 10 ha of pasture

land, and 18 ha of forest. The tributary flows longitudinally

through the wetland, which has a length of 134 m and a

maximum width of 36 m giving a total area of 0.44 ha. The

actual water depth is between 0.3 and 1 m. The first 30 m of

the wetland are free of vegetation, and the remaining area is

covered mainly with Typha capensis Rohrb. (60 percent

coverage), Juncus kraussii Hochst (10 percent coverage),

and Cyperus dives Delile (5 percent coverage).

Sampling Program and Analysis

Two sampling stations were used: inlet sampling was done

in the tributary about 5 m before its entry into the wetland;

outlet sampling took place in the tributary about 7 m below

the wetland. Standard sampling procedure included

measurement of discharge, suspended particles, and

nutrients as well as taking of water and suspended

sediments for pesticide analysis. Discharge was calculated

on the basis of standard formulas (Maniak 1992) using

velocity measurements along cross-section profiles. Turbidity

was measured with a turbidity meter (Dr. Lange) allowing

measurements between 1 and 200 NTU. To calibrate the

turbidity measurements as described by Gippel (1995),

certain samples were filtered through preweighed Whatman

GF/F (0.45 mm pore-size) glass microfibre filters and dried

at 60 °C for 48 hours. The filter paper was then reweighed to

determine total suspended solids (TSS). Ortho-phosphate

and nitrate were measured with photometric test kits from

Merck, Ingelheim. Discharge, TSS, and nutrients were

measured during periods with high-flow conditions between

December 1998 and June 1999 (n = 17 measurements).

Suspended sediments were taken to describe particleassociated pesticide contamination following runoff events.

Time-integrating samples were obtained from continuously

operating suspended particle samplers (Liess and others

1996) installed in the stream bottom. Suspended sediments

were taken out about every 2 weeks between December 14,

1998, (10:00 h; before rainfall event) and May 17, 1999, and

were analyzed for pesticides.

Samples of water for pesticide analysis were collected at the

inlet and outlet site in 3-L glass jars during spray application

on January 11, 28, and 29, 1999. A previous experiment with

the introduction of a tracer (Rhodamin B) 200 m above the

wetland (spraydrift deposition site) was undertaken on

January 7 during the same hydrological conditions

(discharge: 0.026 m3

 per second) to determine the starting

point and duration of sampling intervals at the inlet and

outlet station of the wetland used for the pesticide sampling.

As a result, a 1-hour composite sampling starting 15

minutes after introduction of spraydrift (200 m above the

wetland) was done at the inlet station, whereas 5-hour

composite sampling was performed at the outlet station

starting 4.5 hours after introduction of spraydrift. This

sampling design has been found to enable an integrated

sample of more than 85 percent of the pesticide load to be

obtained at both stations. Water samples (500 to 900 ml)

were solid-phase extracted (SPE) within 10 hours after

sampling using C18 columns (Chromabond) that had been

previously prepared with 6 ml methanol and then 6 ml water.

The columns were air-dried for 30 minutes and kept at –18

°C until analyzed.

Analyses were performed at the Forensic Chemistry

Laboratory of the Department of National Health, Cape

Town. Pesticide residues were extracted from suspended

sediment samples using methanol. Extracts were passed

through a C18 column (Chromabond) and eluted with 2 ml

hexane and then 2 ml dichloromethane. Measurements were

done by means of gas chromatographs (HP 5890) with

electron-capture, nitrogen-phosphorus, and flamephotometric detectors. Water samples were eluted from SPE

columns and measured as well in a GC. The following

detection limits were obtained for water and suspended

sediments: 0.01 μg/l and 0.1 μg per kilogram dw.

Exposure Bioassays

Bloodworms (Chironomus spec.) were used as a test

organism. Animals were obtained from a clean water pond at

Somerset West water treatment plants. The organisms were

collected 1 hour before the exposure started. At each site,

four plastic beakers (9 cm in diameter by 13 cm in height)

containing 20 4th-instar larvae were installed in the stream.

The front and rear walls of the beakers were made of netting

(0.5 mm mesh) to allow water to flow through (current:

0.11±0.02 m per second). In each box, 5 g of mud provided

material for constructing the larval tubes. Mortality was

measured after a 24-hour exposure period. Two trials were

performed, one during a day without any spray application

from December 12, 10:00 hour, to December 13, 10:00 hour,

and one during azinphos-methyl application resulting in
156

spraydrift into the tributary (December 28, 10:00 hour, to

December 29, 10:00 hour).

RESULTS

Sediments and Nutrients

Results obtained for inlet-outlet-measurements of

suspended sediments and nutrients are summarized for

high-discharge conditions in table 1. Trapping efficiency was

calculated by dividing the concentration difference (inlet

minus outlet) by the inlet concentration and expressing it as

a percentage. TSS, ortho-phosphate, and nitrate levels were

reduced by the constructed wetland by 78, 75, and 84

percent. Despite the high inlet concentrations, absolute

values for all three parameters at the outlet station were

quite low, indicating the effectiveness of the wetland.

Particle-Associated Pesticides

The concentrations of azinphos-methyl, chlorpyrifos, and

prothiofos in suspended particles sampled at the inlet station

during periods with rainfall-induced runoff events are

summarized in table 2. Azinphos-methyl concentration in the

inlet sample was 43.3 μg per kilogram during an interval with

a high-rainfall event in December. During the same interval,

prothiofos was present in the inlet sample at a level of 6 μg

per kilogram. Chlorpyrifos occurred at a relatively low

concentration of 0.6 μg per kilogram following a small rainfall

event in February but at elevated concentrations up to 31.4

μg per kilogram during April and May rainfall events. None of

the three OP insecticides were detectable in the outlet

samples at any time, indicating a 100-percent retention of

particle-associated pesticides that were introduced into the

tributary via runoff.

Water-Diluted Pesticides

The effectiveness of retention of water-diluted pesticide was

estimated during spraydrift events following application of

azinphos-methyl (table 3). Inlet concentrations varied during

the three trials between 0.36 and 0.69 μg per liter in the 1-

hour composite samples. Five-hour composite samples

taken at the outlet contained levels between 0.02 and 0.03

μg per liter. Average concentrations detected in the

composite samples and discharge (0.026 m3 per second)

were used to calculate chemical loads for both stations.

Retention was calculated by dividing the load difference

(inlet minus outlet) by the inlet load and expressing it as a

percentage. Retention ranged between 44 and 60 percent

and was, on average, 51 percent.

Exposure Bioassay

Mortality of bloodworms exposed for 24 hours in situ at the

inlet and outlet of the constructed wetland is shown in table

4. During periods without spraying activity, the average

mortality was = 1.25 percent. In the 24-hour period during

spraydrift trial 1, the mortality at the inlet station was

elevated considerably to 41.3 percent, whereas the mortality

at the outlet station stayed low at 2.5 percent.

DISCUSSION

Sediment and Nutrient Trapping Efficiency

As expected, considerable reductions of sediment and

nutrient loads were observed in the wetland. Retention of

TSS was 78 percent during the wet period, when, on

average, 105 mg per liter TSS entered the wetland. Cooper

and Knight (1991) reported a 78-percent trapping efficiency

of a 1.1-ha detention reservoir during storms with inflow TSS

concentrations of 800 mg per liter or greater. A 2.1-ha

Table 1—Mean (n = 17 measurements, ± SE) total

suspended solids and nutrient concentrations of water

at the inlet and outlet of the constructed wetland as well

as trapping efficiency during high discharge conditions

(0.175 ± 0.03 m3

 per second) typical of rainfall conditions

between 2 and 35 mm per day

Trapping

Inlet Outlet efficiency

- - - - - - - - mg/L - - - - - - - Percent

TSS 105 ±14 23.0 ±1.9 78.1

Ortho-phosphate .88± .15 .22± .02 75

Nitrate 1.84± .37 .3 ± .03 83.7

TSS = total suspended solids.

Table 2—Rainfall and concentrations of different OP insecticides associated with suspended particles that were

continously sampled above the constructed wetland during intervals with rainfall-induced runoff events (none of the

suspended particles sampled during the same intervals at the outlet station contained any detectable pesticide levels)

Rainfall characteristics Contamination of inlet station samples

Suspended particlesDate Intensity sampling interval Substance Concentration

 mm/d μg/kg

12/14/98 18.4 12/14–12/29/98 Azinphos-methyl 43.3

02/26/99 2 02/15–03/01/99 Chlorpyrifos .6

04/19/99 18.8 03/31–04/19/99 Chlorpyrifos 31.4

04/21/99 29.8 04/19–05/17/99 Chlorpyrifos 23.9

12/14/98 18.4 12/14–12/29/98 Prothiofos 6
157

constructed wetland receiving constant volumes of nonpointsource agricultural and urban pollution reduced suspended

solids with inlet concentrations of 50 mg per liter by 78

percent and those of 123 mg per liter by 95 percent (Kadlec

and Hey 1994). Lower trapping efficiency observed in the

present study may be due to the small wetland size and the

fact that the wetland is a flow-through wetland without any

water storage capacity and with water inflow rates that vary

over time between 0.016- and 0.522-m3 per second.

Trapping efficiencies during wet conditions for orthophosphate (0.9 mg per liter) and for nitrate (1.8 mg per liter)

were 75 and 84 percent. Similar results (72 percent and 82

percent) have been found in the 1.1-ha Morris Pond, MS

(Cooper and Knight 1990) and in Des Plaines Wetlands, IL,

where total P reduction averaged 74 percent and nitrate

reduction 78 percent (Kadlec and Hey 1994).

The TSS and nutrient-trapping efficiency is still quite high,

although the wetland has never been dug out during the 7-

year period since it was built. This indicates the sustainability

of constructed wetland use as a strategy for the long-term

improvement of water quality.

Retention of Particle-Associated Pesticides: Runoff

Scenario

The constructed wetland acted as a very effective sink for

sediment-associated pesticides during the study period.

Azinphos-methyl (43 μg per kilogram), chlorpyrifos (31 μg

per kilogram), and prothiofos (6 μg per kilogram) detected in

the suspended particles of the inlet water were never

present in the outlet suspended particles. Although it was

thought that wetlands have a high potential to reduce

specifically sorbed chemical load (Rodgers and Dunn 1993),

this fact had not previously been demonstrated for

insecticide fate in small constructed wetlands. It is likely that

one major reason for the effectiveness in retaining particleassociated chemicals is the sedimentation of suspended

particles in the wetland areas with reduced-flow conditions.

As discussed above, TSS were removed at a rate of 78

percent during wet weather conditions. However, the 100-

percent retention detected for all of the three

organophosphate pesticides cannot be explained just by

sedimentation processes. Dilution, decomposition, and

microbial metabolism may account for the remaining

chemical retention (Rodgers and others 1999). The

effectiveness of pesticide retention in wetlands may differ

with season due to changes in water temperature and flow

as well as wetland abiotic and biotic conditions (Spongberg

and Martin-Hayden 1997). There is still a need for further

studies to demonstrate the long-term fate of insecticide

runoff in constructed wetlands.

Retention of Water-Diluted Pesticides:

Spraydrift Scenario

Levels of azinphos-methyl during spraydrift were between

0.36 and 0.69 μg per liter in the inflow water, whereas

maximal concentrations in the outflow were at about 0.07 μg

per liter. It follows that the retention for this pesticide entering

the wetland in the water phase was as high as 51 percent.

There are no other studies dealing with the retention of

water-diluted insecticide input in constructed wetlands.

However, the implementation of retention ponds in

agricultural watersheds was examined as one strategy to

reduce the amount and toxicity of runoff-related insecticide

pollution discharging into estuaries (Scott and others 1999).

For atrazine, a removal rate between 25 and 95 percent was

demonstrated in the Des Plaines Wetland cells (Kadlec and

Hey 1994). Reduction of atrazine concentration in water was

11 to 14 percent in 230-m flow-through wetland mesocosms

(Detenbeck and others 1996). Processes important for

removal of nonpoint-source pesticide runoff in wetlands may

include adsorption, decomposition, and microbial

metabolism (Rodgers and others 1999). The macrophytes

present in the wetland may play an important role in

providing an enlarged surface area for sorption as well as

microbial activity and due to chemical metabolism within the

plants (Wetzel 1993).

Toxicological Evaluation

In situ exposure of bloodworms during spray deposition

clearly indicated an increased 24-hour mortality at the inlet

station in comparison to an only slightly elevated mortality at

the outlet station. It follows that the constructed wetland

makes an important contribution in that it reduces the toxicity

of azinphos-methyl in the tributary water by 94 percent before

it enters the Lourens River. That is, I attribute the observed

reduction of toxicity to the wetland and conclude that it would

be much lower if the tributary water were to flow over the

same distance (230 m) via a conventional channel into the

Lourens River. Azinphos-methyl concentrations measured in

Table 3—Concentration of azinphos-methyl in composite

samples taken during spraydrift at the inlet and outlet of

the constructed wetland. Inlet samples represent 1-hour

composite samples, whereas outlet samples represent

5-hour composite samples (retention was calculated

based on calculation of chemical load)

Event Inlet Outlet Retention

- - - - - - - μg/L - - - - - - - Percent

1 0.69 0.07 49

2 .36 .04 44

3 .62 .05 60

Table 4—Mortality (± SE; n = 4) of Chironomus spec.

exposed for 24 hours at the inlet and outlet of the

constructed wetland during a period without any

spraying activity and during an azinphos-methyl

spraydrift event

Inlet Outlet

Without spraying (%) 1.25 ± 1.25 0

During spraydrift (%) 41.3 ± 2.4 2.5 ± 1.4
158

the inlet water were in the range of acutely toxic

concentrations for various crustacean species including

Gammarus fasciatus Say (Sanders 1972) and Hyalella

azteca Saussure (Ankley and Collyard 1995). They exceeded

the 96-hour LC50 of Chironomus tentans Fabricius, which is

0.37 μg per liter (Ankley and Collyard 1995). According to

prevous experiments, the 24-hour LC50 for the chironomid

species used in the present study is approx. 7.3 μg per liter.

On the basis of the 24-hour LC50 and the azinphos-methyl

concentration measured in the field, it is not possible to

explain the observed mortality of exposed bloodworms.

Similar differences between toxicological reactions in the

field and those predicted from laboratory toxicity tests at the

same contaminant levels have been described by other

authors as well. Matthiesen and others (1995) observed 100

percent mortality of caged Gammarus pulex L. following

exposure to a peak concentration of 27 μg per liter

carbofuran, which exceeded the 24 hour LC50 of 21 μg per

liter only for a period of 3 to 5 hours. Baughman and others

(1989) expected differences in measured and real exposure

concentrations to be a reason for higher mortalities in in situ

bioassays than predicted from laboratory data.

CONCLUSION

It was shown that the investigated wetland is capable of

reducing the load and toxicity of nonpoint agricultural

pesticide pollution. Keeping in mind the relative importance

of runoff-related pesticide contamination associated with

suspended particles, constructed wetlands can be regarded

as a very effective tool for the reduction of nonpoint

agricultural pesticide contamination in small streams.

ACKNOWLEDGMENTS

This study is part of a joint research project between the

University of Stellenbosch, South Africa (Professor A.J.

Reinecke) and the Technical University of Braunschweig,

Germany, and is funded by the National Research

Foundation, Pretoria. The author thanks Corlie Hugo and

Victor Krause for cooperating in the field technical

operations, Sue Peall for performing the pesticide analysis,

and Beate Helling and Ann Thorson for helpful comments on

the manuscript.

REFERENCES

Ankley, G.T.; Collyard, S.A. 1995. Influence of piperonyl butoxide on

the toxicity of organophosphate insecticides to three species of

freshwater benthic invertebrates. Comparative Biochemistry and

Physiology. 110: 149–155.

Baker, L.A. 1993. Introduction to nonpoint source pollution and

wetland mitigation. In: Olson, R.K. comps. Created and natural

wetlands for controlling nonpoint source pollution. Boca Raton,

FL: CRC Press: 12–45.

Baughman, D.S.; Moore, D.W.; Scott, G.I. 1989. A comparison and

evaluation of field and laboratory toxicity tests with fenvalerate

on an estuarine crustacean. Environmental Toxicology and

Chemistry. 8: 417–429.

Cooper, C.M. 1993. Biological effects of agriculturally derived

surface-water pollutants on aquatic systems—a review. Journal

of Environmental Quality. 22: 402–408.

Cooper, C.M.; Knight, S.S. 1990. Nutrient trapping efficiency of a

small detention reservoir. Agricultural Water Management. 18:

149–158.

Cooper, C.M.; Knight, S.S. 1991. Nutrient and sediment trapping

efficiency in small farm ponds. Proceedings for the Environmentally Sound Agriculture Conference: 673–680.

Detenbeck, N.E.; Hernanutz, R.; Allen, K.; Swift, M.C. 1996. Fate

and effects of the herbicide atrazine in flow-through wetland

mesocosms. Environmental Toxicology and Chemistry. 15: 937–

946.

Gippel, C.J. 1995. Potential of turbidity monitoring for measuring the

transport of suspended solids in streams. Hydrological Processes. 9: 83–97.

Groenendijk, P.; van der Kolk, J.W.H.; Travis, K.Z. 1994. Prediction of

exposure concentration in surface waters. In: Hill, I.R.;

Heimbach, F.; Leeuwangh, P.; Matthiessen, P., eds. Freshwater

field tests for hazard assessment of chemicals. Boca Raton, FL:

Lewis Publishers: 105–125.

Hammer, D.A. 1992. Designing constructed wetlands systems to

treat agricultural nonpoint source pollution. Journal of Ecological

Engineering. 1: 49–81.

Kadlec, R.H.; Hey, D.L. 1994. Constucted wetlands for river water

quality improvement. Water Science Technology. 29: 159–168.

Lewis, M.A.; Mayer, F.L.; Powell, R.L. [and others], eds. 1999.

Ecotoxicology and risk assessment for wetlands. Pensacola:

Society of Environmental Toxicology and Chemistry (SETAC).

392 p.

Liess, M.; Schulz, R.; Neumann, M. 1996. A method for monitoring

pesticides bound to suspended particles in small streams.

Chemosphere. 32: 1963–1969.

Loague, K.; Corwin, D.L.; Ellsworth, T.R. 1998. The challenge of

predicting nonpoint source pollution. Environmental Science &

Technology. 32: 130–133.

Mainstone, C.P.; Schofield, K. 1996. Agricultural management for

nonpoint pollution control, with particular refernce to the UK.

European Water Pollution Control. 6: 21–29.

Maniak, U. 1992. Hydrologie und Wasserwirtschaft. Berlin: Springer

Verlag. 568 p.

Matthiesen, P.; Sheahan, D.; Harrison, R. [and others]. 1995. Use of

a Gammarus pulex bioassay to measure the effects of transient

carbofuran runoff from farmland. Ecotoxicology and Environmental Safety. 30: 111–119.

Mitsch, W.J. 1992. Landscape design and the role of created,

restored, and natural riparian wetlands in controlling nonpoint

source pollution. Ecological Engineering. 1: 27–47.

Rodgers, J.H., Jr.; Dickson, G.W.; Dillon, T. [and others]. 1999.

Workgroup V synopsis: constructed wetlands as a risk mitigation

alternative. In: Lewis, M.A.; Mayer, F.L.; Powell, R.L. [and others],

eds. Exotoxicology and risk assessment for wetlands.

Pensacola, FL: Society of Environmental Toxicology and

Chemistry (SETAC): 315–342.

Rodgers, J.H., Jr.; Dunn, A. 1993. Developing design guidelines for

constructed wetlands to remove pesticides from agricultural

runoff. In: Olson, R.K., comps. Created and natural wetlands for

controlling nonpoint source pollution. Boca Raton, FL: CRC

Press: 142–164.
159

Sanders, H.O. 1972. Toxicity of some insecticides to four species of

malacostracan crustaceans. Tech. Pap. 66. Washington, DC: U.S.

Department of the Interior, Fishery and Wildlife Service. 19 p.

Scott, G.I.; Fulton, M.H.; Moore, D.W. [and others]. 1999. Assessment of risk reduction strategies for the management of

agricultural nonpoint source pesticide runoff in estuarine

ecosystem. Toxicology and Industrial Health. 15: 200–213.

Spongberg, A.L.; Martin-Hayden, J.M. 1997. Pesticide stratification

in an engineered wetlands delta. Environmental Science and

Technology. 31: 3161–3165.

Tharme, R.; Ratcliffe, G.; Day, E. 1997. An assessment of the

present ecological condition of the Lourens River, Western

Cape, with particular reference to proposals for stormwater

management. Cape Town, South Africa: Freshwater Research

Unit, University of Capetown. 289 p.

Van der Valk, A.G.; Jolly, R.W. 1993. Recommendation for research

to develop guidelines for the use of wetlands to control rural

nonpoint source pollution. In: Olson, R.K., comps. Created and

natural wetlands for controlling nonpoint source pollution. Boca

Raton, FL: CRC Press: 211–234.

Wetzel, R.G. 1993. Constructed wetlands: scientific foundations are

critical. In: Moshiri, G.A., comp. Constructed wetlands for water

quality improvement. Boca Raton, FL: CRC Press: 3–7.

Wolverton, B.C. 1987. Artificial marshes. In: Reedy, K.R.; Smith,

W.H., comps. Aquatic plants for water treatment and resource

recovery. Orlando, FL: Magnolia Publishing.

Wood, A. 1990. The application of artificial wetlands in South Africa.

In: Cooper, P.F.; Findlater, B.C., comps. Constructed wetlands in

water pollution control. Oxford: Pergamon Press: 235–244.
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INTRODUCTION

Forest cover has declined globally, from an estimated

6.1 billion ha of original forest extent to the present 3.45

billion ha (Krishnaswamy and Hanson 1999). The greatest

loss in cover has occurred in Asia-Pacific, Africa, and

Europe (all more than 60-percent loss of forest cover).

Losses in North America are relatively low (25 percent),

while Latin America (Central and South) has lost over 30

percent of the original forest cover. Nevertheless, the area in

forest plantations is only 135 million ha, although increasing

(Kanowski 1997).

Forest restoration on land cleared for agriculture, often

termed afforestation, is widespread. Land may have been

abandoned because of infertility, frequent flooding, or other

site limitations. Today, as in the past, forest cover in

populated areas is in dynamic equilibrium with land cleared

for agriculture and taken for urban uses. Market forces,

changing trade policies, or agricultural reforms drive

conversion of cleared land back to trees. In Europe, for

example, afforestation is a policy instrument to retire land

from agriculture because of attempts by the European

Union to reduce agricultural subsidies (Madsen and others

2001).

The objectives of this paper are to place the afforestation

efforts in the Lower Mississippi Alluvial Valley (LMAV) into a

global context by drawing parallels to work in other

countries. The challenges of forest restoration in different

countries are surprisingly similar (Kanowski 1997):

overcoming site degradation/limitations, prescribing

appropriate species, and applying cost-effective

establishment methods. Plantation forestry is the most

effective approach to restoration of large areas, and

plantations that are more complex are recent trends.

TERMINOLOGY

What constitutes restoration can be confusing as the term is

used indiscriminately. Changes in land cover and land use

influence the dynamic relationship between degrading and

restoring processes. If we consider the undisturbed,

idealized natural mature forest as a starting point (fig. 1),

then conversions to other land uses such as agriculture or

pasture are through deforestation. Relatively frequent but

moderate disturbance, such as plowing, herbicides, and

grazing, maintain the nonforest cover.

Similarly, a change in both land cover and land use occurs

when forests are converted to urban uses, flooded by dams,

or removed along with topsoil/overburden in mining and

extractive activities. Such drastic conversion usually involves

severe disturbance and is maintained more or less

permanently by structures more than by cultural activities

(fig. 1).

Even-aged harvesting of mature forest in a sustainable

manner is a change of land cover but not land use. A new,

young forest will result from natural regeneration or by

reforestation, i.e., planting trees in a cutover. Unsustainable

harvesting such as high-grading degrades stand structure

or diversity. Pollutant loading, outbreaks of insects or

diseases (especially exotics), invasion by aggressive exotic

plants, or disasters such as hurricanes or wildfires can also

degrade forests. In all these instances, intervention to

restore species diversity or stand structure can be termed

rehabilitation (fig. 1).

Given sufficient time and the cessation of disturbances,

agricultural land, as well as urbanized land, will revert to

forest, if that is the potential natural vegetation as set by

climate. Abandonment and reversion to forests, albeit

FOREST RESTORATION IN A GLOBAL CONTEXT

John A. Stanturf1

1 Project Leader, Disturbance and Management of Southern Ecosystems, USDA Forest Service, 320 Green Street, Athens, GA 30602.

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability of

wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.

Abstract—Forest restoration on land cleared for agriculture is occurring around the world. Often land was abandoned

because of infertility, frequent flooding, or other site limitations. In some countries, market forces or changing trade policies

drive conversion of cleared land to plantations of exotic or native tree species. The objective of this paper is to introduce the

special session on restoration of bottomland hardwoods by placing efforts in the Lower Mississippi Alluvial Valley into a

global context. The challenges of forest restoration are surprisingly similar: overcoming site degradation, prescribing

appropriate species, and applying cost-effective establishment methods. While plantation forestry remains the most

effective approach to large-scale restoration, the trend is toward plantations that are more complex. This trend is characterized by more intimate association with other land uses, more diverse goals for species composition and vegetation

structure in restoration planting, and more direct involvement by landowners in both the conception and implementation of

restoration schemes. Benefits of restoration planting include reduced soil erosion; improved water quality; increased wildlife

habitat; and increased supply of wood for fuel, lumber, and fiber. Increasingly, objectives of restoration planting include

carbon sequestration.
161

secondary or even degraded forest types, will occur

gradually on a time scale of a few decades to centuries.

Human intervention, however, can accelerate the reversion

process. Afforestation of agricultural land may consist simply

of planting trees, although techniques that are more

intensive are available. Reclamation of urbanized land

usually requires more extensive modification, which may

include stabilization of spoil banks or removal of water

control structures, followed by tree planting. Because severe

degradation may limit the possibilities for reclamation, this

process is sometimes called replacement (Bradshaw 1997).

Generally, restoration connotes some transition from a

degraded state to a former natural condition. All the

restorative activities described (reforestation, rehabilitation,

afforestation, and reclamation) have been called forest

restoration, although to the purist none would qualify as true

restoration (Bradshaw 1997, Harrington 1999). In the

narrowest interpretation, restoration requires a return to an

ideal ecosystem with the same previous species diversity,

composition, and structure (Bradshaw 1997) and, as such, is

probably impossible to attain (Cairns 1986). Pragmatically,

forest restoration can describe situations where forest land

use and land cover are restored through the approaches of

afforestation or reclamation.

COMMON CHALLENGES

Three steps are key to planning forest restoration: (1)

understanding current conditions (as a starting point), (2)

clarifying objectives and identifying an appropriate goal, and

(3) defining feasible actions that will move toward the

desired condition. In most cases, the silviculturist may

choose among multiple silvicultural pathways toward the

desired future condition. The choice of intervention affects

the financial cost, the nature of intermediate conditions, and

the time it takes to achieve the desired condition. It is

imperative to make silvicultural decisions with clear

objectives and with an understanding of a particular

intervention’s probable success.

Many examples of forest restoration can be classified as

afforestation, reclamation, or rehabilitation (table 1). Three

papers in this proceedings discuss in detail: bottomland

hardwoods in the Southern United States (Gardiner and

others 2001); broadleaves in the Nordic countries (Madsen

and others 2001); and mangroves in Southeast Asia

(Burbridge and Hallin 2001). The challenges of forest

restoration are surprisingly similar: overcoming site

degradation, prescribing appropriate species, and applying

cost-effective establishment methods.

Overcoming Site Degradation

Previous land use may have degraded site conditions,

especially for afforestation and reclamation projects. The

specific conditions may vary from soil erosion or salinization,

in which physical structure and chemistry of the soil are

inhospitable to native trees, to lowered fertility from

continuous cropping, e.g., Whalley 1988. In some cases,

land becomes available for restoration because it is too

infertile for agriculture. Existing forest stands in need of

rehabilitation may have become degraded by past

mismanagement, such as high-grading (removing all the

biggest and best trees, leaving undesirable species or trees

of poor form or low vigor), fire suppression, or holding water

late into the growing season in green tree reservoirs. In other

cases, hydroperiod alterations, hurricanes, severe

windstorms, floods, or insect outbreaks may have degraded

the stands, but not usually the site.

Mature Forest Mature Forest

Agriculture

Urban

Permanent Water

Resource Extraction

Degraded Degraded 

Forest Young 

Deforestation/ Deforestation/ Forest

Conversion

Conversion

Deforestation

Reclamation

Reclamation Reclamation Afforestation

Regeneration/ Regeneration/

Reforestation

Rehabilitation

Harvest

Pollutants

Insect/Disease

Disasters

Harvest

Figure 1—The terminology of forest restoration is best viewed in terms of land use as well

as land cover change.
162

Site potential and extent of degradation set limits on what

can be achieved by intervention. Site potential refers to the

combination of relatively unchanging physical factors that

affect species composition and stand vigor. Soil and

landform characteristics determine moisture availability,

aeration, and fertility. In wetland forests, hydroperiod

characteristics are important (flood frequency, seasonality,

duration, and depth). Site potential is not immutable,

however, and can be influenced positively or negatively by

changes in land cover or land use.

The cause and possible continuing problem of site or stand

degradation should be identified. For example, alteration of a

site by changed hydroperiod poses several questions. Can

the hydroperiod be restored or the effects of alteration

somehow mitigated? Should the restoration effort target a

vegetation assemblage adapted to present hydroperiod and

site conditions? Hydroperiod alterations caused by flood

control projects, dams, or highway construction tend to be

irreversible, at least in the short term. Flooding caused by

beaver dams, however, can be reduced by removing the

dam, but continued management of beaver population levels

will be required to avoid recurring problems. The guiding

principle should be to rehabilitate or restore in accordance

with existing conditions, unless alteration is feasible,

affordable, and within the control of the silviculturist.

Appropriate Species

Most restoration efforts favor native species, although some

situations may use exotic species. In the Tropics, population

pressures and land scarcity may require that restoration

include species that provide early economic returns

(Grainger 1988, Islam and others 1999, Parrotta 1992), and

native forest species may be unsuited for degraded sites.

Fast-growing exotic species alter site conditions enough for

native species to thrive (Ohta 1990, Parrotta and others

1997). Nevertheless, lack of knowledge may lead to neglect

of native species (Butterfield and Fisher 1994, Fisher 1995,

Knowles and Parrotta 1995).

The perception of native species may be contentious. Some

fast-growing species may be native but considered

undesirable. Public citizens or government agencies may be

averse to planting more pine [especially loblolly (Pinus

taeda L.)] rather than broadleaves in the Southern United

States, or eastern cottonwood (Populus deltoids Bartr. ex

Marsh.) in the LMAV. Species under consideration may be

native to the area but not to the site. In the LMAV, for

example, extensive hydrologic changes have allowed

planting of more oak trees (Quercus spp.) than were

probably in the forests prior to European settlement (King

and Keeland 1999). Even documenting the composition of

the predisturbance forested landscape can be difficult and

contentious (Hamel and Buckner 1998; Stanturf and others,

in press).

An open question is to what extent should the manager

today consider the possible effects of global climate change

in choosing appropriate species to plant. The different

global circulation models used by policymakers yield very

different results for the Southern United States at the scale

of the forest stand. Nevertheless, managers contemplating

long rotations may want to hedge their bets by planting

species adapted to drier conditions on upland sites. In

bottomlands, the situation is more complicated. Rising sea

level will not only inundate coastal forests but also cause a

rise in the base level of rivers in the region, changing the

hydroperiod of many sites. In some bottomlands, species

tolerant of drier conditions would be warranted, but in those

likely to be inundated later in the growing season, it is better

to plant species that are adapted to prolonged soil

saturation.

Table 1—Examples of forest restoration efforts in various parts of the world

Type of restoration Region Former condition Restored condition

Afforestation Lower Mississippi Alluvial

Valley, USA Agriculture Bottomland hardwoods

Afforestation Nordic countries Agriculture Hardwoods, sometimes Norway

spruce

Afforestation Tropical countries Agriculture Exotic and native hardwoods

Afforestation Venezuela Cerrado Caribbean pine

Afforestation Iceland Eroded grazing land Birch, lupine/birch

Reclamation Everywhere Mined land Various

Reclamation Asia Shrimp ponds Mangrove

Reclamation Ireland Mined peatland Sitka spruce, various hardwoods

Reclamation India Saline and sodic soils Eucalyptus spp., Acacia spp.,

other native spp.

Rehabilitation Southeastern United States Loblolly pine plantations Longleaf pine woodlands

Rehabilitation Interior Highlands, Southeastern

United States Shortleaf pine/hardwood Shortleaf pine/bluestem grass

forests woodlands

Rehabilitation Central Europe Norway spruce plantations Oak or beech woodlands

Rehabilitation England and Scotland Spruce or pine plantations Mixed woodlands
163

Effective Establishment Methods

Choosing species appropriate to the site and management

objectives of the landowner is an important first step. Choice

of stock type and proper handling are important, as well as

adequate preparation of the site and postplanting practices,

such as weed control. High survival is needed to insure

adequate stocking (seedling density) and to minimize costs,

especially where seedling costs are high, as in Scandinavia

(Madsen and others 2001). Survival rates in industrial

plantations, commonly 80 to 90 percent, set the benchmark.

However, expecting such high survival in many restoration

programs may be unreasonable, as the knowledge base

may be insufficient due to limited research, practical

experience, or available labor.

Vigorous growth of established seedlings is important to

reach target stocking levels. Low-vigor seedlings may

survive but are at greater risk of mortality from weed

competition, mammal herbivory, or from insects and

diseases. Vigorous growth will also speed the development

of stand structure and canopy closure, important for

achieving conservation and wildlife benefits (Stanturf and

others, in press). On the other hand, cultural practices raise

establishment costs and may not have a lasting effect on

vigor (Stanturf and others 1998).

Planting density is an important decision because of the

effect it has on meeting landowner objectives and minimizing

costs. In order to determine an adequate stocking level given

seedling survival, a simple approach is to calculate the initial

density needed to achieve a future density. For example, the

Wetlands Reserve Program (WRP) is a Federal incentive

program in aid of farmers planting hardwoods on low-lying

cropland (Stanturf and others 2000). The WRP target at age

3 is 309 stems per hectare, which is low for timber

production and probably inadequate for wildlife (Stanturf and

others, in press). Nevertheless, the agency will only share

the cost for planting 750 seedlings per hectare with the

landowner. Therefore, the initial stocking must allow for

intensity of site preparation, planting efficiency, and species

survival rates. Nuttall oak (Q. nuttallii Palmer) is the most

commonly planted oak species in the LMAV and has an

average operational survival rate of 60 percent for planted

seedlings with minimum site preparation. For other oak

species, however, survival is typically lower, 30 to 40

percent; and the target will not be met. In addition,

inexperienced crews plant most WRP sites; and survival

rates are below the operational benchmark, resulting in

significant failures (Stanturf and others, in press).

PLANTATION FORESTRY AS A

RESTORATION MECHANISM

The first step in restoring a forest is to establish trees, the

dominant vegetation. Although this is not full restoration in

the sense of Bradshaw (1997), this necessary step is far

from a trivial accomplishment (Stanturf and others 1998;

Stanturf and others, in press). Nevertheless, many people

object to traditional plantations on the grounds of aesthetics

or lack of stand and landscape diversity. The correct

ecological comparison, however, is between plantations and

intensive agriculture rather than between plantations and a

mature natural forest (Stanturf and others, in press). All

forest alternatives provide at least some vertical structure,

increased plant diversity, and some wildlife and

environmental benefits. Kanowski (1997) argued for a

dichotomy in concepts between the traditional plantations

organized for fiber production and more complex plantation

forests that seek to maximize social benefits other than

wood. Complex plantations that retain the economic and

logistic advantages of simple plantations can meet

restoration goals.

Advantages of Simple Plantations

Simple plantations are single purpose, usually even-aged

monocultures that can produce as much as 10 times greater

wood volume than natural forests (Kanowski 1997). Simple

plantations, nevertheless, provide multiple benefits when

compared to alternatives such as continuous agriculture; if

managed well, they satisfy sustainability criteria. Significant

advantages of simple plantations are that they easily can be

established using proven technology, their management is

straightforward, and they benefit from considerable

economies of scale. If financial return is the primary objective

of a landowner, simple plantations may be preferred and

some restoration goals will be attained (Stanturf and others,

in press). Nevertheless, complex plantations can be

established that provide greater social benefit but at lower

rates of return from timber production, possibly as little as 10

percent less (Kanowski 1997), or even at a net financial gain

to the landowner, e.g., Stanturf and Portwood 1999).

Characteristics of Complex Plantations

Association with other land uses—Objections to

plantations are often cast in terms of aesthetics. The sharp

boundary between a plantation and other land uses is

objectionable, as is the uniformity of trees planted in rows. In

order to integrate the plantation with other land uses, the

sharp edges of plantations can be softened by fuzzy or

curved boundaries. Where plantations are on small farm

holdings, agroforestry systems of intercropping can blend

land uses. Forested riparian buffers are established in

agricultural fields to protect water quality by filtering

sediment, nutrients, and farm chemicals; and they bar easy

access by livestock to stream banks. Riparian buffers add

diversity to the landscape and serve as wildlife corridors

between patches of fragmented forests. In floodplain

landscapes such as bottomland hardwoods, areas of

permanently saturated or inundated soil (respectively, moist

soil units and open water areas) are common and diversify

the interior of plantations.

Several options are available to overcome the uniformity of

rows. Perhaps the simplest technique is to offset the rows.

Uniform spacing between rows and between seedlings

within a row is common, resulting in a square pattern. Rows

can be offset to produce a parallelogram instead of a square.

Alternatively, plantations can be planned with a recreational

viewer in mind so that the view from trails and roads is

always oblique to the rows, thereby escaping notice. At any

rate, once the canopy reaches sufficient height that ground

flora and midstory plants can establish, most plantations

take on the appearance of natural stands, at least to the

casual observer.

Species composition and vegetation structure—A more

serious objection to plantations is the lack of diversity, in
164

terms of species composition and vertical structure.

Essentially, simple plantations are not as diverse as natural

stands, at least for many years. Foresters have devised

several methods to establish multiple species stands. For

example, planting several blocks of different species in a

stand, or even alternate rows of different species is possible

and creates some diversity at the stand level. Distribution,

however, remains more clumped than would be typical of a

natural stand.

Other methods are available, including nurse crops of faster

growing native species (Schweitzer and others 1997) or

exotics (Ashton and others 1997, Lamb and Tomlinson

1994). In this approach, there is no intention of retaining the

nurse crop species throughout the rotation of the slower

growing species. (This could also be termed relay

intercropping.) While the nurse-crop method has many

advantages and in the short term provides species diversity

and probably vertical structure, once the nurse crop is

removed the residual stand may lack diversity. The challenge

is to develop methods for establishing several species in

intimate mixtures, such as would occur in a natural stand,

but avoiding excessive mortality during the self-thinning or

stem exclusion stage of stand development. Such methods

must account for the growth patterns of the species, relative

shade tolerances, and competitive ability.

Vertical structure is an important feature of forests for wildlife

(DeGraaf 1987; Hamel and others, in press; Twedt and

Portwood 1997). Early stages of stand development,

whether in natural forests or plantations, are characterized

by low light in the understory until crowns differentiate. In

most restoration forests, little development of the understory

and midstory occurs for many years. Annual disturbance

while in agriculture removed buried seed and rootstocks of

native plants, and low light levels in the young forest

preclude understory development from invaders. To

accommodate this deficiency, the manager can intervene to

plant understory species; at present, little research affords

guidance on methods, planting density, or probable success

rates. As indicated above, relay intercropping provides

vertical structure for a time. Natural dispersal into gaps can

also encourage understory development, whether gaps are

created by thinning or left during planting (Allen 1997,

Otsamo 2000). The critical factor limiting understory

development by natural invasion is whether there are seed

sources for the understory plants within dispersal range

(Chapman and Chapman 1999, Johnson 1988).

BENEFITS OF RESTORATION

The benefits of restoration are usually identified in terms of

agency priorities or social benefits; seldom are the diverse

objectives of landowners recognized. In most market

economies where rights and obligations of ownership rest

with private landowners, what is appropriate for public land

may not be the most attractive restoration option for private

landowners (Stanturf and others, in press). Nevertheless,

there can be considerable overlap in the expected benefits

to society and the affected landowner. The array of possible

landowner objectives can be illustrated with a limited set of

management scenarios from the LMAV (table 2). For

simplification, three scenarios are presented: (1) shortrotation management for pulpwood or fuelwood; (2) a longer

rotation typical of management for saw log production which

is suitable for wildlife species that require complex vertical

structure, such as certain neotropical migratory songbirds

(Hamel and others, in press); and (3) an option termed

green vegetation, which is essentially the no-management

scenario. In the green vegetation scenario, species

composition and stand structure are secondary concerns to

removing land from active agriculture. This option meets the

objectives of Federal programs, such as the WRP (Stanturf

and others, in press). It may also provide habitat conditions

for certain wildlife that otherwise would not occur on the

landscape (Hamel and others, in press).

Table 2—Financial, recreational, and environmental benefits expected from three afforestation scenarios common in the

Lower Mississippi Alluvial Valley, Southern United States

Expected benefit level

Financial Recreational Environmental

Scenario Short term Long term Hunting Nonconsumptive Conservation practices Land retirement

Short rotation

(pulpwood,

fuelwood) High High High Medium Medium No

Long rotation

(timber,

wildlife) Medium High High High High Medium

Green

vegetation Low to no No Low Medium Medium High
165

Benefits are comprised of financial, recreational, and

environmental outcomes. Because cash flow is important to

many landowners, and the adjustment from annual to

periodic income is often cited as a barrier to afforestation,

financial benefits are considered as both short term and

long term. Recreational benefits are hunting (typically for

deer, turkey, and waterfowl) and nonconsumptive benefits,

such as bird watching or hiking. Environmental benefits are

separated into conservation practices, such as those

installed to control soil erosion and protect water quality or

enhance wildlife habitat, and land retirement, where there is

no on-going management activity.

Financial Benefits

Financial returns from active management are substantial

relative to the green vegetation scenario. Saw log rotations

of high-value oak and green ash (Fraxinus pennsylvanica

Marsh.) are expected within 60 to 80 years, with the first

commercial thinning beginning in 20 to 30 years. Short-term

financial returns from growing pulpwood-sized eastern

cottonwood in the LMAV are realized within 10 years of

afforestation (Stanturf and Portwood 1999). The short-term

financial returns are low from plantations of other species.

Nevertheless, other species can be combined with

cottonwood in the nurse-crop technique to produce income

for one or two pulpwood rotations, hence the medium rating.

The green vegetation scenario, typified by the WRP

plantings, provides no long-term income, as timber

management is unlikely given the understocked stands that

develop (Stanturf and others, in press). In the short term,

there is income from the one-time easement payment made

to the landowner (Stanturf and others 2000).

Other income can be realized by some landowners from

hunting leases and potentially from carbon sequestration

payments. In the Mississippi portion of the LMAV, hunting

rights are leased for $7.50 to $12.35 per hectare per year.

There is also a potential for considerable income to

landowners from credits from carbon sequestration (Barker

and others 1996). While there is considerable uncertainty

over the accounting for carbon credits under the Kyoto

Protocol, there seems to be agreement that afforestation will

be eligible for offset credit (Schlamadinger and Marland

2000). Current projections in the United States for the value

of a carbon credit are on the order of $2.72 to $4.54 per

megagram of CO2

 sequestered, but the value is much higher

in Europe. In Norway, for example, there is already a carbon

tax on gasoline equivalent to $49 per megagram of CO2

(Solberg 1997). Estimates from economic models suggest

that a carbon tax of $27 to $109 per megagram of CO2 would

be necessary to stabilize global emissions at the 1990 level

(Solberg 1997). Under these conditions, growing biomass for

fuel would become an attractive alternative to fossil fuel

because biofuels have no net impact on global carbon

levels. At some time in the future, landowners in the LMAV

may want to optimize carbon sequestration and biofuel

benefits by planting willow (Salix spp.) on soils too wet for

cottonwood.

Recreational Benefits

The primary recreational benefits assumed in the examples

are from creating and enhancing wildlife habitat. Not all

wildlife species require the same kind of habitat, so for

simplicity the expected benefits can be separated into

recreational hunting by the landowner (rather than lease

fees) and nonconsumptive wildlife activities, such as bird

watching or simply the existence value of wildlife to the

landowner. Most species hunted in the LMAV benefit from a

range of forest conditions and expected benefits are high in

stands managed for pulpwood or saw logs. Low expected

value is derived from the kind of open stands likely to

develop from the green vegetation scenario (Allen 1997,

King and Keeland 1999). Neotropical migratory birds and

other birds are not uniform in their habitat requirements

(Hamel and others, in press) but some will benefit from the

kind of early successional habitat typical of short-rotation

stands (Twedt and Portwood 1997), as well as the early

successional herbaceous fields of the green vegetation

scenario. Species of concern are of two kinds, those

requiring the early successional herbaceous vegetation and

those found in the kind of complex vegetation structure

found only in older stands, which the saw log rotation may

develop in time (Hamel and others, in press). Birds that use

the intermediate conditions of stand development are

probably likely to occur in developing stands for which the

intended management purpose is sawtimber production.

Environmental Benefits

Water-quality benefits of afforestation accrue from reducing

soil erosion (Joslin and Schoenholtz 1998), and filtering,

retaining, and assimilating nutrients and farm chemicals

from surface runoff and groundwater (Huang and others

1990). As Lockaby and Stanturf (2001) point out, however,

typical restoration stands in the LMAV no longer experience

the kind of flow-through hydrology of a riverine system; and

the filtering action will be limited.

Greater water-quality benefit will be derived from forested

riparian buffers. Planted forested buffer strips in an

agricultural landscape are uncommon, although several

studies have examined the filtering action of natural forested

riparian zones (Cooper and Gilliam 1987; Cooper and others

1987; Lowrance and others 1983, 1984a, 1984b, 1986;

Peterjohn and Correll 1984; Todd and others 1983).

Comerford and others (1992) summarized these studies and

concluded that buffer strips are quite effective in removing

soluble nitrogen and phosphorus (up to 99 percent) and

sediment. The efficiency of pesticide removal by forested

buffer strips has been examined in some environmental fate

studies that concluded that buffer strips 15 m or wider were

generally effective in minimizing pesticide contamination of

streams from overland flow (Comerford and others 1992).

Recently, forested buffer strips in the LMAV became

attractive financially to the landowner by a new incentive

program (Continuous Signup/Conservation Reserve

Program), which allows use of the cottonwood/red oak

nurse-crop system.

CONCLUSION

Forest restoration, in the broad sense that encompasses

afforestation, rehabilitation, and reclamation, is occurring

throughout the temperate and boreal zones. Site conditions

differ, native species are diverse, and the policy context in

which restoration occurs varies. Nevertheless, the

challenges faced by managers are similar: overcome site

degradation, prescribe appropriate species, and apply cost-
166

effective methods. Clarity of objectives is critical to designing

a successful restoration program and diagnosis of site

conditions and potential should guide intervention.

Knowledge and experience with establishing plantation for

timber production can be used to efficiently restore large

areas of agricultural land to forest.

While simple plantations have many financial and

technological advantages, plantations that are more complex

will be required in most countries. Aesthetics, species

diversity, and the need to rapidly create vertical vegetation

structure are some of the concerns that must be addressed.

The benefits of restoration should be viewed in comparison

to the previous conditions of land cover and land use. In

most cases, environmental benefits are immediate.

Restoration forests can differ in functioning and

management, according to the objectives of the landowner.

There is usually considerable overlap between social and

individual benefits. Even forests primarily managed for

timber production provide environmental and recreational

benefits to society.

LITERATURE CITED

Allen, J.A. 1997. Reforestation of bottomland hardwoods and the

issue of woody species diversity. Restoration Ecology. 5: 125–

134.

Ashton, P.M.S.; Gamage, S.; Gunatilleke, I.A.U.N.; Gunatilleke, C.V.C.

1997. Restoration of a Sri Lankan rainforest: using Caribbean

pine Pinus caribaea as a nurse for establishing late-successional

tree species. Journal of Applied Ecology. 34: 915–925.

Barker, J.R.; Baumgardner, G.A.; Turner, D.P.; Lee, J.J. 1996. Carbon

dynamics of the conservation and wetland reserve programs.

Journal of Soil and Water Conservation. 51: 340–346.

Bradshaw, A.D. 1997. What do we mean by restoration? In:

Urbanska, K.M.; Webb, N.R.; Edwards, P.J., eds. Restoration

ecology and sustainable development. Cambridge, UK: Cambridge University Press: 8–14.

Burbridge, P.R.; Hallin, D.C. 2002. Rehabilitation of coastal wetland

forests degraded through their conversion to shrimp farms.

[This proceedings].

Butterfield, R.P.; Fisher, R.F. 1994. Untapped potential: native

species for reforestation. Journal of Forestry. 92 (6): 37–40.

Cairns, J., Jr. 1986. Restoration, reclamation, and regeneration of

degraded or destroyed ecosystems. In: Soule, M.E., ed. Conservation biology. Ann Arbor, MI: Sinauer Publishers: 465–484.

Chapman, C.A.; Chapman, L.J. 1999. Forest restoration in abandoned agricultural land: a case study from East Africa. Conservation Biology. 13: 1301–1311.

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. NCASI

Tech. Bull. 631. New York: National Council of the Paper Industry

for Air and Stream Improvement. 48 p.

Cooper, J.R.; Gilliam, J.W. 1987. Phosphorus redistribution from

cultivated fields into riparian areas. Soil Science Society of

America Journal. 51: 1600–1604.

Cooper, J.R.; Gilliam, J.W.; Daniels, R.B. 1987. Riparian areas as

filters for agricultural sediment. Soil Science Society of America

Journal. 51: 416–420.

DeGraaf, R.M. 1987. Managing northern hardwoods for breeding

birds. In: Nyland, R.D., ed. Managing northern hardwoods,

Proceedings silvicultural symposium; 1986 June 23–25;

Syracuse, NY. Misc. Publ. 13 (ESF 87–002). Syracuse, NY:

Syracuse University, Faculty of Forestry: 348–362.

Fisher, R.F. 1995. Amelioration of degraded rain forest soils by

plantations of native trees. Soil Science Society of America

Journal. 59: 544–549.

Gardiner, E.S.; Russell, D.R.; Oliver, M., Dorris, L.C., Jr. 2001.

Bottomland hardwood afforestation: state of the art. In: Holland,

M.J.; Warren, M.L., Jr.; Stanturf, J.A.

Grainger, A. 1988. Estimating areas of degraded tropical lands

requiring replenishment of forest cover. International Tree Crops

Journal. 5: 31–61.

Hamel, P.B.; Buckner, E.R. 1998. How far could a squirrel travel in

the treetops? A prehistory of the southern forest. In:

Wadsworth, K.G., ed. Transactions of the 63rd North American

wildlife and natural resources conference; 1998 March 20–24;

Orlando, FL. Washington, DC: Wildlife Management Institute:

309–315.

Hamel, P.B.; Nuttle, T.; Woodson, C.A.; Broerman, F. 2002. Forest

restoration as ecological succession: should we speed it up or

slow it down? [This proceedings].

Harrington, C.A. 1999. Forests planted for ecosystem restoration or

conservation. New Forests. 17 (1–3): 175–190.

Huang, W.-Y.; Algozin, K.; Ervin, D.; Hickenbotham, T. 1990. Using

the conservation reserve program to protect groundwater quality.

Journal of Soil and Water Conservation. 46: 251–254.

Islam, K.R.; Kamaluddin, M.; Bhuiyan, M.K.; Badruddin, B. 1999.

Comparative performance of exotic and indigenous forest

species for tropical semievergreen degraded forestland reforestation in Chittagong, Bangladesh: Land Degradation & Development. 10: 241–249.

Johnson, W.C. 1988. Estimating dispersibility of Acer, Fraxinus, and

Tilia in fragmented landscapes from patterns of seedling

establishment. Landscape Ecology. 1: 175–187.

Joslin, J.D.; Schoenholtz, S.H. 1998. Measuring the environmental

effects of converting cropland to short-rotation woody crops: A

research approach. Biomass and Bioenergy. 15: 301–311.

Kanowski, P.J. 1997. Afforestation and plantation forestry. In:

Proceedings of the XI World Forestry Congress; 1997 October

13–22; Antalya, Turkey. [Available on the Web: http://

coombs.anu.edu.au/Depts/RSPAS/RMAP/kanow.htm].

King, S.L.; Keeland, B.D. 1999. Evaluation of reforestation in the

Lower Mississippi River Alluvial Valley. Restoration Ecology. 7:

348–359.

Knowles, O.H.; Parrotta, J.A. 1995. Amazonian forest restoration: an

innovative system for native species selection based on

phenological data and field performance indices. Commonwealth

Forestry Review. 74 (3): 230–243.

Krishnaswamy, A.; Hanson, A., ed. 1999. Our forests, our future:

summary report, world commission on forests and sustainable

development. Cambridge, UK: Cambridge University Press: 37.
167

Lamb, D.; Tomlinson, M. 1994. Forest rehabilitation in the AsiaPacific region; past lessons and present uncertainties. Journal

Tropical Forest Science. 7 (1): 157–170.

Lockaby, B.G.; Stanturf, J.A. 2002. Potential effects of restoration

on biogeochemical functions of bottomland hardwood ecosystems. [This proceedings].

Lowrance, R.R.; Todd, R.L.; Asmussen, L.E. 1983. Waterborne

nutrient budgets for the riparian zone of an agricultural watershed. Agriculture Ecosystems and Environment. 10: 371–384.

Lowrance, R.R.; Todd, R.L.; Asmussen, L.E. 1984a. Nutrient cycling

in an agricultural watershed: I. Phreatic movement. Journal of

Environmental Quality. 13: 22–27.

Lowrance, R.R.; Todd, R.L.; Fail, J. Jr. [and others]. 1984b. Riparian

forest as nutrient filters in agricultural watersheds. BioScience.

34: 374–377.

Lowrance, R.R.; Sharpe, J.K.; Sheridan, J.M. 1986. Long-term

sediment deposition in the riparian zone of a Coastal Plain

watershed. Journal of Soil and Water Conservation. 41: 266–271.

Madsen, P.; Arad•ttir, Á.L.; Gardiner, E. [and others]. 2002. Forest

restoration in the Nordic countries. [This proceedings].

Ohta, S. 1990. Initial soil changes associated with afforestation with

Acacia auriculiformis and Pinus kesiya on denuded grasslands

of the Pantabangan area, Central Luzon, the Philippines. Soil

Science and Plant Nutrition. 36: 633–643.

Otsamo, R. 2000. Early development of three planted indigenous

tree species and natural understory vegetation in artificial gaps

in an Acacia mangium stand on an Imperata cylindrica grassland

site in South Kalimantan, Indonesia. New Forests. 19: 51–68.

Parrotta, J.A. 1992. The role of plantation forests in rehabilitating

tropical ecosystems. Agriculture, Ecosystems, and Environment.

41: 115–133.

Parrotta, J.A.; Turnbull, J.W.; Jones, N. 1997. Catalyzing native

forest regeneration on degraded tropical lands. Forest Ecology

and Management. 99: 1–7.

Peterjohn, W.T.; Correll, D.L. 1984. Nutrient dynamics in an

agricultural watershed: observations on the role of riparian

forests. Ecology. 65: 1466–1475.

Schlamadinger, B.; Marland, G. 2000. Land use & global climate

change—forests, land management, and the Kyoto Protocol.

Arlington, VA: Pew Center on Global Climate Change. 54 p.

Schweitzer, C.J.; Stanturf, J.A.; Shepard, J.P. [and others]. 1997.

Large-scale comparison of reforestation techniques commonly

used in the Lower Mississippi River Alluvial Valley. In: Pallardy,

S.G.; Cecich, R.A.; Garrett, H.G.; Johnson, P.S., eds. Proceedings, 11th central hardwood forest conference; 1997 March 23–

26; Columbia, MO. Gen. Tech. Rep. NC–188. St. Paul, MN: U.S.

Department of Agriculture, Forest Service, North Central Forest

Experiment Station: 313–320.

Solberg, B. 1997. Forest biomass as carbon sink—economic value

and forest management/policy implications. Critical Reviews in

Environmental Science and Technology. 27(Special): S323–

S333.

Stanturf, J.A.; Gardiner, E.S.; Hamel, P.B. [and others]. 2000.

Restoring bottomland hardwood ecosystems in the Lower

Mississippi Alluvial Valley. Journal of Forestry. 98(8): 10–16.

Stanturf, J.A.; Portwood, C.J. 1999. Economics of afforestation with

Eastern cottonwood (Populous deltoides) on agricultural land in

the Lower Mississippi Alluvial Valley. In: Haywood, James D. ed.

Proceedings of the tenth biennial southern silvicultural research

conference; 1999 February 16–18; Shreveport, LA. Gen. Tech.

Rep. SRS–30. Asheville, NC: U.S. Department of Agriculture,

Forest Service, Southern Research Station: 66–72

Stanturf, J.A.; Schoenholtz, S.H.; Schweitzer, C.J.; Shepard, J.P.

2001. Achieving restoration success: myths in bottomland

hardwood forests. Restoration Ecology. 9(2): 189–200.

Stanturf, J.A.; Schweitzer, C.J.; Gardiner, E.S. 1998. Afforestation of

marginal agricultural land in the Lower Mississippi River Alluvial

Valley, USA. Silva Fennica. 32: 281–297.

Todd, R.L.; Lowrance, R.R.; Hendrickson, O. [and others]. 1983.

Riparian vegetation as filters of nutrients exported from a

Coastal Plain agricultural watershed. In Lowrance, R.R.; Todd,

R.L.; Asmussen, L.E.; Leonard, R.A., eds. Nutrient cycling in

agricultural ecosystems. Spec. Publ. 23. Athens, GA: University

of Georgia, College of Agriculture: 485–496.

Twedt, D.J.; Portwood, C.J. 1997. Bottomland hardwood reforestation for neotropical migratory birds: are we missing the forest for

the trees? Wildlife Society Bulletin. 25: 647–652.

Whalley, D.N. 1988. Afforestation of salt-affected soils—an international problem. International Plant Propagator’s Society

Combined Proceedings. 37: 212–221.
168

INTRODUCTION

Large-scale reforestation emphasizing riparian zones can

dramatically affect fish communities of low-gradient Coastal

Plain streams. Forested riparian zones provide multiple

benefits to stream fishes (Angermeier and Karr 1984,

Gregory and others 1991). Indirect benefits include

reduction of sediment and nutrient inputs (Lowrance and

others 1984), stabilization of stream banks, and moderation

of water temperature extremes (Gregory and others 1991).

These factors can affect fish productivity, physiology,

reproduction, and community composition (Matthews 1987).

More directly, organic matter input into streams as leaves

and in-stream wood provides the primary energy source for

aquatic macroinvertebrates (Wallace and others 1997),

which form the food base for most stream fishes. In sandy

Coastal Plain streams, debris dams and large wood greatly

increase macroinvertebrate production (Benke and others

1984, Smock and others 1989), promote channel stability,

and increase habitat complexity for fishes (Shields and

Smith 1992).

From 1949 to 1985 the U.S. Department of Agriculture,

Forest Service, and other Federal agencies led the YazooLittle Tallahatchie Project in the upper Coastal Plain of

northern Mississippi (U.S. Department of Agriculture, Forest

Service 1988). The program was designed to re-establish

forests and stabilize soils in a region subjected to a

sequence of massive erosion cycles from 1830 into the

1970s (Schumm and others 1984). Postsettlement alluvium

from hillside erosion, reaching 5-m depths in valleys,

exacerbated flooding and prompted the dredging and

channelization of most streams in the region (Schumm and

others 1984). Modification of stream channels induced

cycles of stream incision or headcutting, an active

geomorphic process on the landscape today (Shields and

others 1994, 1998). Although not focused on riparian areas,

the reforestation effort successfully stabilized hillslopes and

dramatically increased forest cover of the region.

In 1999 we sampled fish and fish habitat in streams within

the stabilized, reforested region. In these sandy, upper

Coastal Plain streams, we explored the relationships among

in-stream wood, canopy cover, and stream fish

assemblages. We were particularly interested in how

extensive channel modification affected these relationships.

We examined whether multivariate relationships among

canopy, in-stream wood, and fish community attributes were

detectable, and if so, how they covaried across a range of

watershed sizes and local habitat configurations.

STUDY REGION

The study region lies primarily within the Holly Springs

National Forest located in Benton, Lafayette, Marshall,

Tippah, and Union Counties in north-central Mississippi. The

proclamation boundaries of potential Federal ownership

include 153,000 ha, about 35 percent of which is national

forest land. The ecoregion is classified as the Northern

Loessial Hills Subsection, Coastal Plain Middle Section,

Southeastern Mixed Forest Province (Keys and others

1995).

The study streams drain parts of four major river basins: the

Little Tallahatchie and Yocona Rivers (both Yazoo River

tributaries) and the Hatchie and Wolf Rivers (both direct

Mississippi River tributaries). The topography consists of

irregular low hills (200-m maximum relief) dissected by welldeveloped, dendritic drainage systems with flat, often broad,

floodplains adjacent to large streams. Stream substrate is

predominantly sand or silty sand with accumulations of finer

material in depositional areas.

FOREST LINKAGES TO DIVERSITY AND ABUNDANCE

IN LOWLAND STREAM FISH COMMUNITIES

Melvin L. Warren, Jr., Wendell R. Haag, and Susan B. Adams1

1 Warren, Research Biologist, Haag, Fisheries Biologist, and Adams, Research Fisheries Biologist, Center for Bottomland Hardwoods

Research, Southern Research Station, 1000 Front Street, Oxford, MS 38655.

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability of

wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.

Abstract—In 1999 we sampled fish and fish habitat in 79 stream reaches within watersheds of north-central Mississippi.

Despite a program of successful reforestation and soil stabilization (Yazoo-Little Tallahatchie Project, 1949–1985), nearly all

streams in the region are channelized or incised. In these sandy, upper Coastal Plain streams, we explored the relationships

among in-stream wood, canopy cover, and stream fish assemblages and were particularly interested in how these relationships are affected by extensive channel modification. Minnows, sunfishes, darters, and catfishes, respectively, dominated

the fauna. Total fish species, total fish abundance, minnow relative abundance, and canopy cover were related only to

watershed size. Stream incision, as indicated by high banks and shallow water depths, showed negative associations with

sunfish relative abundance, in-stream wood, and detritus. Flow and large in-stream wood were associated positively with

relative abundances of darters and catfishes. Despite scarcity of wood and deep pools in these systems, we detected

associations between fish assemblage composition and in-stream wood. Our analysis suggests even modest densities of

in-stream wood can shift fish assemblage attributes from colonizing stages to intermediate or stable stages.
169

The disturbance history of the area has profoundly modified

stream systems through aggradation, channelization, and

incision. The soils of the region are highly erodible, being

derived from wind-deposited loess and Coastal Plain

materials (Dendy and others 1979). Following European

settlement, the land was deforested and cultivated. These

actions accelerated hillside erosion, led to massive gully

formation in uplands, and filled valleys with postsettlement

alluvium (Cooper and Knight 1991, Schumm and others

1984, Shields and others 1994). All of these disturbances

exacerbated flooding. Between about 1840 and 1930

individuals and local drainage districts attempted

ineffectively to reclaim valley lands through channelization of

streams and construction of drainage ditches (Shields and

others 1994). From 1930 to 1970 in conjunction with

reforestation and soil stabilization programs, Federal

agencies led a second round of stream channelization and

constructed many flood-control impoundments in headwater

streams. Streams responded to changes in channel

gradients, resulting in large part from channelization by

undergoing incision. This process affects entire stream

networks upstream to the smallest headwater channels.

Incision is rapid, proceeding upstream at rates up to 500 m

per year (Schumm and others 1984), often occurs

episodically, and is widespread in the region (Shields and

others 1998). First, streams eroded through the partially

consolidated surface material; then, downcutting proceeded

rapidly through underlying strata of unconsolidated sand and

clay. Geologic controls in the resulting sand-bed channels

are infrequent and mostly consist of outcrops of

consolidated clay or cemented sand (Shields and others

1997). Despite successful reforestation and concomitant soil

stabilization in the area, few streams have escaped

channelization or incision. Massive bank failures from stream

channel incision can contribute amounts of sediment to the

region’s stream systems comparable to amounts produced

by deforestation at European settlement (Schumm and

others 1984).

METHODS

We sampled fish and fish habitat at 79 wadeable stream

reaches in the study region from May to September 1999

(appendix A). These reaches represent a range of

watershed sizes, habitat types, and watershed conditions

across the study region. When possible, we selected one or

more extreme headwater, middle, and downstream reaches

in each watershed and covered as many habitat types as

possible within selected reaches, e.g., riffle, run, pool. We

sampled a predetermined length of stream, such that

sampling effort was approximately proportional to stream

size (Angermeier and Smogor 1995, Lyons 1992, Paller

1995). Typically, reach lengths were 20 times the average

width of the stream. Minimum reach length was 80 m for

streams < 4 m in average width, and maximum reach length

was 300 m for streams wider than 15 m. We subdivided each

reach into four subreaches of equal length. We located the

upstream and downstream termini of the reach and

subreaches with a hip chain, tape, or range finder before

sampling.

Fish Sampling

We employed two methods of fish sampling at each reach:

(1) single-pass backpack electrofishing, and (2) seining.

Combination of the two methods helped reduce bias and

assured capture of most fishes representative of the reach

assemblage. We standardized effort for both methods and

sampled each subreach separately. For electrofishing, we

calculated total sampling time as the reach length times 5

seconds and applied one-fourth the total effort to each

subreach. For seining, we conducted 8 seine hauls (2 per

subreach) for streams < 5 m in average width and 12 seine

hauls (3 per subreach) for streams > 5 m in average width.

One haul was either a sustained drag of the seine within a

stream macrohabitat, such as a pool or one set-and-kick

seine in a riffle (Jenkins and Burkhead 1994).

Standardization of fishing times for electrofishing and effort

for seining assured that streams of all sizes were sampled

with effort proportional to their size. We kept fish samples

from each subreach separately for both electrofishing and

seining, but pooled all data within each reach for this

analysis.

Physical Habitat Sampling

Subsequent to fish sampling, we sampled physical habitat.

We established 12 equally spaced transects within each

reach so that 3 transects occurred (one-fourth, one-half, and

three-fourths of subreach length) within each subreach. At

each transect, we recorded wetted width and estimated

bank stability (eroding or stable); bank angle (steep, 90

degrees; moderately steep, > 45 degrees; gradual, > 45

degrees); bank height (nearest meter); and dominant bank

vegetation. At equally spaced points along each transect, we

measured water depth and velocity (at 0.6 depth). We

visually categorized dominant substrate type as soft clay,

hard clay, silt, sand, fine gravel (2 to 15 mm), coarse gravel

(16 to 63 mm), cobble (64 to 256 mm), or bedrock. We also

recorded presence or absence of detritus, small wood (< 10

cm in diameter or < 1.5 m in length), large wood (>10 cm in

diameter, > 1.5 m in length), and aquatic vegetation within

an area approximately 1 m2

 around each point. We

estimated canopy cover at each point as 0, 25, 50, or 100

percent cover. We adjusted the number of points per transect

for stream width. Transects > 10 m in width had points

located at 2-m intervals; transects 5 to 10 m in width, at 1-m

intervals; and transects < 5 m had a minimum of 5 equally

placed points.

Data Analysis

We calculated the following measures of fish community

composition for 66 of 79 reaches: total species, total

abundance, and familial relative abundance. We eliminated

13 reaches from the analysis because physical data were

incomplete, no fishes were captured, or visits in September

1999 revealed the sampled reach was located in an

intermittent stream. Total species and total abundance at a

reach were the total recorded number of fish species and

individuals, respectively. Familial relative abundance was the

proportional abundance of the four most abundant families

across reaches: Centrarchidae, sunfishes; Cyprinidae,

minnows; Percidae, darters (genera Etheostoma and

Percina); and Ictaluridae, catfishes (mostly bullhead

catfishes, Ameiurus spp., and madtom catfishes, Noturus

spp.). Familial responses to physical habitat differences are

ecological in origin, and within each family, fish species are

reasonably similar in habitat requirements (Shields and

others 1998).
170

For each reach, we derived mean values for a suite of

physical and riparian characteristics. We calculated average

depth (centimeter), velocity (meter per second), and bank

height (meter) from transect points. We determined

proportions of detritus, small in-stream wood, and large instream wood from the number of points at which each

occurred divided by the number of points surveyed. For

comparison with other studies, we also were interested in

converting the large in-stream wood data into the number of

pieces per unit length. Assuming that each transect point

with large in-stream wood represented one piece of large

wood, we calculated the number of pieces over the length of

each reach and averaged across reaches. This rough data

transformation to pieces of wood per unit length allows

general comparison with other studies. We determined

watershed area on 1:100,000-scale topographic maps in a

Geographic Information System. We calculated stream order

(Strahler 1957) and link magnitude (Osborne and Wiley

1992), including both intermittent and permanent channels

on 1:24,000-scale topographic maps.

We used principal components analysis (PCA) as a first

approximation to partition covariation of physical variables

and fish community composition into one or more

component axes and to ordinate reaches in physical

variable-fish community space. The PCA contained

geomorphic-related variables (watershed area, average

velocity, average depth, and average bank height); riparianrelated variables (proportions of canopy cover, large wood,

small wood, and detritus); and fish community variables

(total species, total abundance, and familial proportional

abundance). We extracted principal components (SAS

1996) from the correlation matrix of the variables. We

transformed watershed area, total species, average depth,

and average velocity to logarithms (base 10) and

proportional variables to arcsine square roots (Sokal and

Rohlf 1995). We used the broken-stick model (Jackson

1993) to evaluate the interpretability of each principal

component axis. Using this method, an axis has interpretive

value if observed eigenvalues exceed eigenvalues

generated by the model.

RESULTS

The sampled reaches represented a range of physical

stream conditions (table 1). Watershed area spanned 3

orders of magnitude; however, 55 of 66 reaches drained

watersheds of 100 to 10 000 ha. Stream order ranged from

first to fifth order and link magnitude from 1 to 314, but 82

percent of reaches were third order. Streams were generally

shallow with low velocities. Only seven reaches had mean

depths > 30 cm and only eight had velocities > 0.30 m per

second. Stream incision as indicated by average bank

height and percent shallow depths was a prominent feature

across reaches. Most reaches (66 percent) had bank

heights > 2 m, notwithstanding inclusion of many small

streams. Most depths (61 percent) along transects were <

15 cm. Stream substrate was dominated by sand (71.4

percent of all transect points) with some silt (14.3 percent),

hard clay (7.3 percent), and soft clay (5.1 percent); rocky

substrates were rarely encountered (< 2.0 percent). Average

canopy cover was high despite the range in watershed

sizes. Three-fourths of reaches had 50 percent or greater

canopy cover. The proportion of large wood was low

(table 1), and for half the reaches, large wood occurred in <

5 percent of the sampled points. Large wood averaged 47

pieces per kilometer (SE = 5) and ranged from 0 to 213

pieces per kilometer.

The fish fauna was relatively diverse and abundant within

most reaches, but highly variable among reaches (tables 2,

3). We captured 65 fish species representing 15 families

(table 3), and only 16 reaches had fewer than 10 species.

However, variability of total species among reaches was

high with a coefficient of variation (CV) of 46.2 percent (table

2). Total abundance also showed high variability with CV of

99.8 percent. Catch per unit effort for electrofishing ranged

from 0.02 to 0.71 individuals per second (mean = 0.21, SE =

0.016, CV = 62.4 percent). Capture rate on a reach-length

basis ranged from 0.09 to 6.64 individuals per meter (mean

= 1.47, SE = 0.138, CV = 76.4 percent). Density of fishes

ranged from 0.04 to 1.7 individuals per square meter (mean

= O.42, SE = 0.043, CV = 82.5 percent).

Table 1—Mean, standard error (SE), minimum, and

maximum values for watershed area, physical habitat

variables, canopy cover, and in-stream wood variables at

66 reaches in upper Coastal Plain streams of northcentral Mississippi

Variable Mean SE Min. Max.

Watershed area (ha) 2543 457.4 43 15194

Depth (cm) 14.3 1.40 2 49

Velocity (m/s) .17 .013 < .01 .53

Bank height (m) 2.7 .18 .9 7.8

Canopy cover (%) 71 3.3 6 100

Detritus (%) 16 2.0 0 95

Small wood (%) 27 2.2 0 68

Large wood (%) 7 .9 0 28

Table 2—Mean, standard error (SE), minimum, and

maximum values for fish community variables at 66

reaches

Variable Mean SE Min. Max.

Total species 14.7 0.84 3 33

Total abundance

(individuals) 154.2 18.94 7 1009

Minnow relative

abundance (%) 41.7 3.41 0 100

Sunfish relative

abundance (%) 21.0 2.50 0 100

Darter relative

abundance (%) 13.3 1.80 0 73

Catfish relative

abundance (%) 6.3 .99 0 32
171

Table 3—Fish species frequency and proportional occurrence for 66

stream reaches in the upper Coastal Plain of north-central Mississippi

Reaches

Family and

species Proportion Frequency

Amiidae

Bowfin

(Amia calva Linnaeus) .015 1

Aphredoderidae

Pirate perch

[Aphredoderus sayanus (Gilliams)] .288 19

Atherinidae

Brook silverside

[Labidesthes sicculus (Cope)] .121 8

Catostomidae

Creek chubsucker

[Erimyzon oblongus (Mitchill)] .364 24

Northern hog sucker

[Hypentelium nigricans (Lesueur)] .212 14

Spotted sucker

[Minytrema melanops (Rafinesque)] .045 3

Blacktail redhorse

(Moxostoma poecilurum Jordan) .182 12

Centrarchidae

Green sunfish

(Lepomis cyanellus Rafinesque) .606 40

Warmouth

[L. gulosus (Cuvier)] .273 18

Orangespotted sunfish

[L. humilis (Girard)] .030 2

Bluegill

(L. macrochirus Rafinesque) .758 50

Dollar sunfish

[L. marginatus (Holbrook)] .258 17

Longear sunfish

[L. megalotis (Rafinesque)] .545 36

Redear sunfish

[L. microlophus (Gunther)] .045 3

Redspotted sunfish

[L. miniatus (Jordan)] .121 8

Spotted bass

[Micropterus punctulatus (Rafinesque)] .288 19

Largemouth bass

[M. salmoides (Lacepede)] .303 20

White crappie

(Pomoxis annularis Rafinesque) .030 2

Clupeidae

Gizzard shad

[Dorosoma cepedianum (Lesueur)] .015 1

Cyprinidae

Bluntface shiner

[Cyprinella camura (Jordan & Meek)] .606 40

Blacktail shiner

(C. venusta Girard) .303 20

continued
172

Table 3—Fish species frequency and proportional occurrence for 66

stream reaches in the upper Coastal Plain of north-central Mississippi

(continued)

Reaches

Family and

species Proportion Frequency

Cyprinidae (continued)

Common carp

(Cyprinus carpio Linnaeus) .015 1

Cypress minnow

(Hybognathus hayi Jordan) .015 1

Mississippi silvery minnow

(Hybognathus nuchalis Agassiz) .242 16

Striped shiner

(Luxilus chrysocephalus Rafinesque) .197 13

Ribbon shiner

[Lythrurus fumeus (Evermann)] .136 9

Redfin shiner

[L. umbratilis (Girard)] .409 27

Golden shiner

[Notemigonus crysoleucas (Mitchill)] .106 7

Orangefin shiner

(Notropis ammophilus Suttkus & Boschung) .030 2

Emerald shiner

(N. atherinoides Rafinesque) .288 19

Yazoo shiner

(N. rafinesquei Suttkus) .258 17

Mimic shiner

[Notropis volucellus (Cope)] .121 8

Pugnose minnow

(Opsopoeodus emiliae Hay) .045 3

Bluntnose minnow

[Pimephales notatus (Rafinesque)] .288 19

Bullhead minnow

[P. vigilax (Baird & Girard)] .061 4

Creek chub

[Semotilus atromaculatus (Mitchill)] .530 35

Esocidae

Grass pickerel

(Esox americanus Gmelin) .121 8

Fundulidae

Blackstripe topminnow

[Fundulus notatus (Rafinesque)] .561 37

Blackspotted topminnow

[F. olivaceous (Storer)] .845 56

Ictaluridae

Yellow bullhead

[Ameiurus natalis (Lesueur)] .333 22

Brown bullhead

[A. nebulosus (Lesueur)] .030 2

Channel catfish

[Ictalurus punctatus (Rafinesque)] .159 11

Least madtom

[Noturus hildebrandi (Bailey & Taylor)] .015 1

Brindled madtom

(N. miurus Jordan) .121 8

continued
173

Table 3—Fish species frequency and proportional occurrence for 66

stream reaches in the upper Coastal Plain of north-central Mississippi

(continued)

Reaches

Family and

species Proportion Frequency

Ictaluridae (continued)

Brown madtom

(N. phaeus Taylor) .667 44

Flathead catfish

[Pylodictis olivaris (Rafinesque)] .015 1

Lepisosteidae

Spotted gar

[Lepisosteus oculatus (Winchell)] .076 5

Longnose gar

[L. osseus (Linnaeus)] .015 1

Percidae

Bluntnose darter

[Etheostoma chlorosoma (Hay)] .045 3

Slough darter

[E. gracile (Girard)] .167 11

Harlequin darter

(E. histrio Jordan & Gilbert) .045 3

Brighteye darter

[E. lynceum (Hay)] .333 22

Johnny darter

(E. nigrum Rafinesque) .258 17

Goldstripe darter

(E. parvipinne Gilbert & Swain) .424 28

Cypress darter

[E. proeliare (Hay)] .106 7

Yazoo darter

(E. raneyi Suttkus & Bart) .409 27

Speckled darter

[E. stigmaeum (Jordan)] .303 2

Gulf darter

[E. swaini (Jordan)] .212 14

Redfin darter

[E. whipplei (Girard)] .242 16

Bandfin darter

(E. zonistium Bailey & Etnier) .015 1

Dusky darter

[Percina sciera (Swain)] .515 34

River darter

[P. shumardi (Girard)] .030 2

Petromyzontidae

Ichthyomyzon sp. .424 28

Poeciliidae

Western mosquitofish

[Gambusia affinis (Baird & Girard)] .273 18

Sciaenidae

Freshwater drum

(Aplodinotus grunniens Rafinesque) .061 4
174

Minnows, sunfishes, darters, and catfishes, respectively,

dominated the fauna and on average accounted for about 82

percent of individuals captured (table 2). These four families

were represented by at least one species in most reaches

(table 3). Minnow and sunfish species each occurred in 65

reaches, darter species in 63, and catfish species in 55.

Average relative abundance and number of species showed

similar ranks among families. Minnows had the highest

average relative abundance and the most species (17) and

catfishes, the lowest abundance and 7 species (table 3).

Sunfishes and darters were represented by 11 and 14

species, respectively. The 11 other families accounted for 16

species that ranged from relatively rare, e.g. bowfin, gizzard

shad, to widespread, e.g., blackspotted and blackstripe

topminnows (table 3).

Principal components analysis of physical and fish

community variables across reaches produced eigenvalues

exceeding those of the broken-stick model, providing

support for interpretation of the first three axes (table 4). The

first three axes accounted for 60 percent of the total variance

(table 4). We considered variables with loadings > 0.20

(absolute value) as being associated with an axis; 52

percent of loadings across axes did not exceed this value.

Because we were interested in interaction of stream

geomorphic variables with in-stream wood and fish

community variables, we named axes based on the

geomorphic variable(s) with the highest absolute loading.

We interpreted the first principal component axis (PC–I) as a

watershed size axis. Magnitudes and polarities of loadings

(table 4) contrasted total species (fig. 1A), total abundance

(fig. 1B), minnow relative abundance, watershed area, depth,

and velocity, which all increased with watershed size, with

canopy cover and detritus accumulations, which decreased

with watershed size. Total species, total abundance, minnow

relative abundance, and canopy cover were only associated

with PC–I; absolute loadings on other axes were low. We

detected no strong associations between canopy cover and

these fish community variables except in the context of

watershed size. The watershed size axis was only weakly

associated with bank height, an indicator of stream incision.

We interpreted PC–II as a stream incision-depth axis.

Magnitudes and polarities of loadings contrasted sunfish

abundance, large and small in-stream wood, detritus, and

depth (pool development) with bank height (table 4). The

axis arrayed reaches along a gradient from shallow, deeply

incised streams with low proportions of wood and sunfishes

to deep streams with low banks and relatively high

proportions of wood and sunfishes.

We interpreted PC–III as a stream velocity axis. The

loadings indicate positive covariation between darter and

catfish abundance, velocity, and large wood. Sunfish

abundance and detritus covaried negatively with these

variables. The axis arrayed reaches along a gradient from

flowing, stream habitats having high proportions of large

wood, darters, and catfishes and low proportions of detritus

and sunfishes to slow or nonflowing habitats with low

proportions of large wood, darters, and catfishes and high

proportions of sunfishes. Importantly, the pattern of

covariation indicates that the relative abundance of two

largely rheophilic families, darters and catfishes, covaried

positively with large wood in the context of flowing habitats.

DISCUSSION

We detected covariation among fish community measures,

canopy, in-stream wood, detritus, and physical habitat

configuration across a variety of watershed sizes and

conditions. The watershed size axis essentially removed

species-area effects from the analysis, such that remaining

interpretable axes were independent of watershed size

effects. Total species, minnow relative abundance, and total

abundance were related positively, and percent canopy

cover, negatively to watershed size. Surprisingly, little or no

variability in minnow relative abundance, total species, or

fish abundance was related to either the stream incisiondepth or stream velocity axes, both of which were

associated with in-stream wood and/or detritus. Large wood

increases habitat heterogeneity (Shields and Smith 1992),

and stable, deep, heterogeneous habitats typically yield

increased fish species richness (Angermeier and Smogor

1995, Schlosser 1987) and show decreased dominance by

minnows (Schlosser 1987). We expected that a high

proportion of in-stream wood, especially large wood in

pools, would be associated with increased diversity of fishes

and decreased dominance of minnows after accounting for

watershed size on PC–I. Given the increases in stream

productivity associated with debris dams and large wood

(Benke and others 1984, Smock and others 1989), we also

expected total fish abundance to covary positively with instream wood. In contrast, we found no relationship between

in-stream wood and total species, minnow relative

abundance, or total abundance.

Influences of pool development and flow regimes on instream wood abundance might explain why total species,

minnow relative abundance, and total abundance were not

Table 4—Loadings on the first three principal

component axes of habitat, canopy, in-stream wood, and

fish community variables for 66 stream reaches in the

upper Coastal Plain of north-central Mississippi

Variable PC–I PC–II PC–III

Watershed area 0.421 0.147 0.049

Depth .304 .391 - .052

Velocity .214 .158 .280

Bank height .165 - .368 .066

Canopy cover - .285 - .053 .066

Detritus - .227 .282 - .243

Small wood - .183 .484 .021

Large wood - .030 .480 .222

Total species .402 .173 - .024

Total abundance .396 .037 - .062

Minnow relative abundance .382 - .141 - .030

Sunfish relative abundance - .075 .235 - .525

Darter relative abundance - .068 .061 .543

Catfish relative abundance - .122 .086 .471

Total variance (%) 29 17 14
175

0

5

10

15

20

25

30

35

12345

Log (watershed area)

Total fish species

1

2

3

1234

Log (watershed area)

Log (no. individuals)

Figure 1—Scatterplots of watershed area and total number of fish species (A) and total number of fishes (B) in 66 stream

reaches of the upper Coastal Plain of northern Mississippi.

A

B
176

associated strongly with in-stream wood. The overall amount

of large wood in the sampled reaches appeared to be low,

relative to conditions prevailing in streams within late

successional forests. Large wood in Coastal Plain streams

draining late successional forests is expected to approach

about 200 pieces per kilometer (Smock and others 1989),

and in eastern upland forests it can reach 400 to 500 pieces

per kilometer (Hedman and others 1996, Hildebrand and

others 1997). Our estimate of large wood across the study

region is < 25 percent of the Coastal Plain expectation and

only 6 reaches (9 percent) exceeded 100 pieces per

kilometer. Shields and others (1995) also indicated scarcity

of in-stream large wood in the study area. Maximum depth,

an indicator of pool development, also was limited in the

study reaches. Only 19 percent of our transect points

exceeded 30 cm in depth, and only 6 percent exceeded 60

cm. Wide channels, shallow depths, and flashy flows in

streams of the study region (Shields and others 1995,

Wallerstein and others 1997) likely limit wood accumulation

even in systems with relatively high wood recruitment. We

believe the paucity of deep pools and large wood precluded

detection of strong associations among total species, fish

abundance, and in-stream wood after accounting for

watershed size. Obversely, many of the minnow species

represented in study reaches are adapted to the shallow,

uniform, sandy habitats where they often occur in schools of

hundreds to thousands of individuals. Ubiquity and extreme

abundance of minnows, when treated as a single group, may

explain their strong association with watershed size and lack

of association with other axes. We believe analysis by

species or habitat guilds within minnows may reveal

associations beyond watershed size that were masked by

examining the family as a group.

Stream incision strongly affects the interaction of in-stream

wood with the quality of fish habitats and the composition of

fish assemblages. Pairing Schlosser’s (1987) conceptual

stream framework for fishes and incised channel models

(Schumm and others 1984, Simon 1989), three stages in

fish community development for incising channels were

hypothesized (Shields and others 1998). Colonizing fish

assemblages are associated with shallow, uniform habitats

with little debris. These are dominated by small species,

such as minnows. In relatively undisturbed watersheds,

these assemblages are found in small, headwater streams,

but channelization and incision produce colonizing

assemblages even in large streams. Intermediate

assemblages typify streams with some increase in pool

volume and begin to be comprised of larger fishes, such as

sunfishes (Shields and others 1998). In both colonizing and

intermediate stages, in-stream wood density may be

depressed in response to changes in channel geometry

(Shields and others 1998). As pool depth and volume

increase further, stable assemblages develop with fewer, but

larger, piscivores. Abundance of small invertivores and

omnivores decreases as predation and resulting competition

for refugia among prey species increases (Schlosser 1987,

Shields and other 1998). At this stage, shallow riffle areas

between pools provide important habitat, e.g., refuge from

predators, for benthic invertivores, e.g., darters, madtom

catfishes. Channelization and cycles of channel incision

produce colonizing to intermediate fish assemblages

(Shields and others 1998), but stable assemblage attributes

may develop between disturbance events because of

channel and adjoining floodplain dynamics.

We detected positive associations between the relative

abundance of a dominant family, the sunfishes, and woody,

deep habitats, a response attenuated by stream incision.

These results are congruent with Schlosser’s (1987)

conceptual framework for small stream fish communities, in

which deep, woody habitat was a key factor, and with fish

assemblage changes observed after habitat manipulation in

incised streams (Shields and others 1998). Addition of

relatively deep, stable pools to incised streams shifted fish

assemblages from colonizing stages to those more

representative of intermediate stages, a change principally

involving sunfishes (Shields and others 1998). Deep, woody

pools increase habitat (pool volume), temporal stability, and

forage availability, which influence fish community structure

through shifts in fish age and size structure, species

composition, and trophic composition. Wood in the deepest

habitats in a stream may be especially critical to sunfishes

during summer low flows, when total habitat area is reduced,

and pools become relatively accessible to bird and mammal

predators (Angermeier and Karr 1984). In Coastal Plain

rivers sunfishes, especially Lepomis spp., obtain a large

proportion of their food intake from invertebrates produced

on wood, forming a distinct trophic pathway from in-stream

wood to sunfishes (Benke and others 1985). Despite incision,

poor pool development, and scarcity of large wood, the

results from the stream incision-depth axis indicate that even

modest increases in in-stream wood and pool development

can play a strong role in structuring fish assemblages.

On the stream velocity axis, we discovered further evidence

of the association of large wood and fish assemblage

composition but in the context of water velocity, not pool

development. The fish assemblage showed increased darter

and catfish relative abundances and decreased sunfish

relative abundance as large wood and velocities increased.

Notably, this assemblage pattern emerged after watershed

size and incision-related variation was removed, suggesting

relatively small changes in habitat configuration improved

conditions for species in both families. Although pool

development is important in fish assemblage composition

(Schlosser 1987), we suggest this pattern reflects the

importance of large wood in providing cover for darters and

catfishes irrespective of watershed size, incision, or the

relatively shallow depths and low flows observed in study

reaches. We attribute the trend in sunfish relative abundance

on this axis, at least in part, to their affinity for low velocities,

but realize other more complex factors, e.g., predation, may

be involved. During field sampling we noticed that the

majority of darters and catfishes (particularly madtoms)

showed high affinity for cover, rarely occurring in open,

uniformly sand-bottom habitats. Large wood generally formed

primary cover in flowing habitats and is likely critical to the

persistence of many darters and madtom catfishes (Chan

and Parsons 2000, Monzyk and others 1997) in the study

area. In sum, our results indicated large wood increases

habitat quality for these two dominant and diverse families.

Fish species occurring in the region are those that have

persisted through extensive historical and ongoing

modifications to stream channels. We believe for most
177

stream reaches surveyed that the number of species was at

or near saturation because the available local species pool

lacks potential new colonists regardless of the availability of

stable, heterogeneous habitats. We did discover fish

assemblage patterns associated with in-stream wood, but

the legacy of channelization and incision limits the amount

and thus the influence of wood on fish habitats. The results

on the watershed size axis suggest that most reaches,

regardless of watershed size, showed attributes of

colonizing assemblages, particularly dominance by

minnows. Minnows dominated 61 percent of our reaches.

Minnows predictably dominate uniform shallow reaches

(Schlosser 1987) and are conspicuously abundant in incised

or channelized streams (Shields and others 1998). In

contrast, covariation on the stream incision-depth axis

indicates some reaches show characteristics of intermediate

assemblages with increased representation of sunfishes.

Incision apparently limits sunfishes by creating unstable

shallow habitats that do not retain wood for cover or pool

formation. The higher relative abundances of darters and

catfishes on the stream velocity axis suggests that

colonizing assemblages, predominated by minnows, may

develop stable attributes if large wood and flow are present

in sufficient quantity.

We did not examine some important aspects of stream fish

assemblages, such as size structure or trophic relationships.

Other assemblage attributes might show stronger

associations with amounts of in-stream wood, canopy cover,

or other aspects of habitat configuration. We plan to examine

the data set to discern further assemblage patterns in the

context of ecological guilds and watershed conditions.

ACKNOWLEDGMENTS

We appreciate the reviews of an earlier draft provided by A.

Dolloff, G. McWhirter, D. Shields, and S. Zarnoch.

Discussions with A. Clingenpeel, D. Marion, S. Ross, and D.

Shields were valuable in guiding data analysis and

interpretation. We thank T. Fletcher, C. Harwell, D. Martinovic,

G. McWhirter, A. Pabst, R. Reekstin, and L. Staton for

conducting the field survey. H. Halverson, C. Harwell, G.

McWhirter, and L. Staton were instrumental in entering,

proofing, and summarizing data. G. Henderson provided

assistance in data management and summaries, and C.

Jenkins and F. McEwen provided administrative assistance.

S. Krieger, C. Kilcrease (both Holly Springs National Forest),

and A. Clingenpeel (Ouachita National Forest) graciously

provided maps, global positioning tools, and other

geographic information on streams in the area. G. Yeck

(Holly Springs National Forest) provided staff time, vehicles,

and other courtesies. We are extremely grateful to numerous

private landowners that granted permission to survey

streams on their property. The study was partially supported

by a U.S. Department of Agriculture Forest Service Chief’s

Grant to the Holly Springs National Forest.

LITERATURE CITED

Angermeier, P.L.; Karr, J.R. 1984. Relationship between woody

debris and fish habitat in a small warmwater stream. Transactions of the American Fisheries Society. 113: 716–726.

Angermeier, P.L.; Smogor, R.A. 1995. Estimating number of species

and relative abundances in stream-fish communities: effects of

sampling effort and discontinuous distributions. Canadian

Journal of Fisheries and Aquatic Sciences. 52: 939–949.

Benke, A.C.; Henry, R.L., III; Gillespie, D.M.; Hunter, R.J. 1985.

Importance of snag habitat for animal production in southeastern

streams. Fisheries. 10(5): 8–13.

Benke, A.C.; Van Arsdall, T.C., Jr.; Gillespie, D.M.; Parrish, F.K. 1984.

Invertebrate productivity in a subtropical blackwater river: the

importance of snag habitat and life history. Ecological Monographs. 54: 25–63.

Chan, M.D.; Parsons, G.R. 2000. Aspects of brown madtom,

Noturus phaeus, life history in northern Mississippi. Copeia.

2000: 757–762.

Cooper, C.M.; Knight, S.S. 1991. Water quality cycles in two hill land

streams subjected to natural, municipal, and non-point agricultural stresses in the Yazoo Basin of Mississippi, USA (1985–

1987). Organ der Internationalen Vereinigung fur Theoretische

und Angewandte Limnologie. 24: 1654–1663.

Dendy, F.E.; Ursic, S.J.; Bowie, A.J. 1979. Sediment sources and

yields from upland watersheds in north Mississippi. In: Proceedings of the Mississippi water resources conference; 1979;

Mississippi State, MS. Mississippi State, MS: Mississippi State

University, Water Resources Research Institute: 49–54.

Gregory, S.V.; Swanson, F.J.; McKee, W.A.; Cummins, K.W. 1991. An

ecosystem perspective of riparian zones. BioScience. 41: 540–

551.

Hedman, C.; Van Lear, D.H.; Swank, W.T. 1996. In-stream large

woody debris loading and riparian forest seral stage associations

in the Southern Appalachian Mountains. Canadian Journal of

Forest Research. 26: 1218–1227.

Hilderbrand, R.H.; Lemly, A.D.; Dolloff, C.A.; Harpster, K.L. 1997.

Effects of large woody debris placement on stream channels and

benthic macroinvertebrates. Canadian Journal of Fisheries and

Aquatic Sciences. 54: 931–939.

Jackson, D.A. 1993. Stopping rules in principal components

analysis: a comparison of heuristic and statistical approaches.

Ecology. 74: 2204–2214.

Jenkins, R.E.; Burkhead, N.M. 1994. Freshwater fishes of Virginia.

Bethesda, MD: American Fisheries Society. 1,079 p.

Keys, J.E., Jr.; Carpenter, C.A.; Hooks, S. [and others]. 1995.

Ecological units of the Eastern United States: first approximation. Atlanta: U.S. Department of Agriculture, Forest Service,

Southern Region. 1: 3.500,000; presentation scale; color. [Map].

Lowrance, T.; Todd, R.; Fail, J., Jr. [and others]. 1984. Riparian

forests as nutrient filters in agricultural watersheds. BioScience.

34: 374–377.

Lyons, J. 1992. The length of stream to sample with a towed

electrofishing unit when fish species richness is estimated. North

American Journal of Fisheries Management. 12: 198–203.

Matthews, W.J. 1987. Physicochemical tolerance and selectivity of

stream fishes as related to their geographic ranges and local

distributions. In: Matthews, W.J.; Heins, D.C., eds. Community

and evolutionary ecology of North American stream fishes.

Norman, OK: University of Oklahoma Press: 111–120.
178

Monzyk, F.R.; Kelso, W.E.; Rutherford, D.A. 1997. Characteristics of

woody cover used by brown madtoms and pirate perch in

Coastal Plain streams. Transactions of the American Fisheries

Society. 126: 665–675.

Osborne, L.L.; Wiley, M.J. 1992. Influence of tributary spatial

position on the structure of warmwater fish communities.

Canadian Journal of Fisheries and Aquatic Sciences. 49: 671–

681.

Paller, M.H. 1995. Relationships among number of fish species

sampled, reach length surveyed, and sampling effort in South

Carolina Coastal Plain streams. North American Journal of

Fisheries Management. 15: 110–120.

Schlosser, I.J. 1987. A conceptual framework for fish communities in

small warmwater streams. In: Matthews, W.J.; Heins, D.C., eds.

Community and evolutionary ecology of North American stream

fishes. Norman, OK: University of Oklahoma Press: 17–24.

Schumm, S.A.; Harvey, M.D.; Watson, C.C. 1984. Incised channels:

morphology, dynamics and control. Littleton, CO: Water

Resources Publications. 200 p.

Shields, F.D., Jr.; Knight, S.S.; Cooper, C.M. 1994. Effects of channel

incision on base flow stream habitats and fishes. Environmental

Management. 18: 43–57.

Shields, F.D., Jr.; Knight, S.S.; Cooper, C.M. 1995. Use of the index

of biotic integrity to assess physical habitat degradation in

warmwater streams. Hydrobiologia. 312: 191–208.

Shields, F.D., Jr.; Knight, S.S.; Cooper, C.M. 1997. Rehabilitation of

warmwater stream ecosystems following channel incision.

Ecological Engineering. 8: 93–116.

Shields, F.D., Jr.; Knight, S.S.; Cooper, C.M. 1998. Rehabilitation of

aquatic habitats in warmwater streams damaged by channel

incision in Mississippi. Hydrobiologia. 382: 63–86.

Shields, F.D., Jr.; Smith, R.H. 1992. Effects of large woody debris

removal on physical characteristics of a sand-bed river. Aquatic

Conservation: Marine and Freshwater Ecosystems. 2: 145–163.

Simon, A. 1989. A model of channel response in disturbed alluvial

streams. Water Resources Bulletin. 25: 1171–1188.

Smock, L.A.; Metzler, G.M.; Gladden, J.E. 1989. Role of debris

dams in the structure and functioning of low-gradient headwater

streams. Ecology. 70: 764–765.

Sokal, R.R.; Rohlf, R.J. 1995. Biometry. New York: W.H. Freeman

and Co. 887 p.

Strahler, A.N. 1957. Quantitative analysis of watershed geomorphology. Transactions of the American Geophysical Union. 38: 913–

920.

SAS Institute Inc. 1996. Statistical analysis software for windows.

Version 6.12. Cary, NC: SAS Institute Inc.

U.S. Department of Agriculture, Forest Service. 1988. The YazooLittle Tallahatchie flood prevention project: a history of the Forest

Service’s role. For. Rep. R8–FR8. Atlanta: U.S. Department of

Agriculture, Forest Service, Southern Region. 63 p.

Wallace, J.B.; Eggert, S.L.; Meyer, J.L.; Webster, J.B. 1997. Multiple

trophic levels of a forest stream linked to terrestrial litter inputs.

Science. 277: 102–104.

Wallerstein, N.; Thorne, C.R.; Doyle, M.W. 1997. Spatial distribution

and impact of large woody debris in northern Mississippi. In:

Wang, S.S.Y.; Langenoen, E.J.; Shields, F.D., Jr., eds. Proceedings of the conference on management of landscapes disturbed

by channel incision; 1997 May 19–23; Oxford, MS. Oxford, MS:

University of Mississippi, Center for Computational Hydroscience

and Engineering: 145–150.

Localities for 79 stream reaches surveyed for fish and fish

habitat in north-central Mississippi. An asterisk indicates the

site was not used in the analyses because physical data

were incomplete, no fishes were captured, or visits in

September 1999 revealed the sampled reach was located in

an intermittent stream. Abbreviations: trib. = tributary; CR =

county road; FS = Forest Service road; quad. = U.S.

Geological Survey 1:24,000 scale topographic map; DMS =

degrees, minutes, and seconds; Lat. = latitude; Long. =

longitude; DD = decimal degrees; T = township; R = range; S

= section; and Q = quarter.

HATCHIE RIVER DRAINAGE

Unnamed trib. N. Br. Hurricane Creek, Muddy Creek trib.,

Tippah Co. Date: 7/26/99. CR 324 end, 3 km SE Camp Hill,

21 km NE Ashland. Walnut quad. Lat. (DMS): N34-56-14.2

Long. (DMS): W88-59-41.0 Lat. (DD): 34.9357 Long. (DD): -

88.990167 T-2S, R-3E, S-4, Q-SW. Stream order: 2. Link

magnitude: 4.

APPENDIX A

Unnamed trib. N. Br. Hurricane Creek, Muddy Creek trib.,

Tippah Co. Date: 7/26/99. FS 644A end, 3 km SE Camp Hill,

20 km NE Ashland. Walnut quad. Lat. (DMS): N34-55-41.8

Long. (DMS): W88-59-35.1 Lat. (DD): 34.923633 Long. (DD):

-88.989183 T-2S, R-3E, S-9, Q-NW. Stream order: 3. Link

magnitude: 14.

S. Br. Hurricane Creek, Muddy Creek trib., Tippah Co. Date:

7/27/99. CR 300 bridge, 8 km NNW Falkner, 7 km SE Camp

Hill. Walnut quad. Lat. (DMS): N34-54-18.5 Long. (DMS):

W88-57-51.4 Lat. (DD): 34.903083 Long. (DD): -88.958567

T-2S, R-3E, S-14, Q-SW. Stream order: 4. Link magnitude:

81.

West Prong Muddy Creek, Muddy Creek trib., Tippah Co.

Date: 7/16/99. CR 414 bridge, 17.6 km E Ashland, 8.6 km

NNW Ripley. Falkner quad. Lat. (DMS): N34-49-28.0 Long.

(DMS): W88-58-59.0 Lat. (DD): 34.821333 Long. (DD): -

88.9765 T-3S, R-3E, S-16, Q-NE. Stream order: 4. Link

magnitude: 27.
179

*Unnamed trib. West Prong Muddy Creek, Muddy Creek

trib., Tippah Co. Date: 7/16/99. CR 346, 15 km E Ashland, 14

km NNW Ripley. Whitten Town quad. Lat. (DMS): N34-50-

55.6 Long. (DMS): W89-01-08.3 Lat. (DD): 34.8426 Long.

(DD): -89.01805 T-3S, R-3E, S-6, Q-SE. Stream order: 1.

Link magnitude: 1.

LITTLE TALLAHATCHIE RIVER DRAINAGE

*Bagley Creek, Lafayette Co. Date: 6/7/99. CR 244

(Riverside Rd.), 8 km E Abbeville, 19 km SW Potts Camp.

Malone quad. Lat. (DMS): N34-30-08.3 Long. (DMS): W89-

24-48.1 Lat. (DD): 34.501383 Long. (DD): -89.408017 T-7S,

R-2W, S-4, Q-S. Stream order: 4. Link magnitude: 17.

Bagley Creek, Lafayette Co. Date: 6/7/99. Bagley Lake

outflow, 8.7 km ESE Abbeville, 20.7 km SSW Potts Camp.

Bagley quad. Lat. (DMS): N34-28-50.7 Long. (DMS): W89-

24-19.5 Lat. (DD): 34.475117 Long. (DD): -89.40325 T-7S, R2W, S-9, Q-SE. Stream order: 3. Link magnitude: 10.

Cypress Creek, Lafayette Co. Date: 6/1/99. CR 251 bridge, 6

km WNW Pleasantdale Church, 4 km SE Union Hill Church.

Puskus quad. Lat. (DMS): N34-23-36.7 Long. (DMS): W89-

17-12.9 Lat. (DD): 34.38945 Long. (DD): -89.285483 T-7S, R1W, S-15, Q-NE. Stream order: 3. Link magnitude: 11.

Cypress Creek, Lafayette Co. Date: 6/4/99. 1.6 km upstream

Goolsby Lake, 20.7 km E Oxford, 5.8 km NNW Lafayette

Springs. Denmark quad. Lat. (DMS): N34-21-35.4 Long.

(DMS): W89-18-15.3 Lat. (DD): 34.3559 Long. (DD): -

89.30255 T-8S, R-1W, S-28, Q-N. Stream order: 2. Link

magnitude: 2.

Cypress Creek, Lafayette Co. Date: 6/4/99. CR 252 bridge, 8

km NNW Lafayette Springs, 23.2 km E Oxford. Puskus quad.

Lat. (DMS): N34-22-56.4 Long. (DMS): W89-17-54.3 Lat.

(DD): 34.376067 Long. (DD): -89.292383 T-8S, R-1W, S-15,

Q-SW. Stream order: 2. Link magnitude: 4.

Cypress Creek, Lafayette Co. Date: 5/28/99. CR 244 bridge,

3.6 km WNW Etta, 6.8 km S Cornersville. Puskus quad. Lat.

(DMS): N34-28-37.7 Long. (DMS): W89-15-45.1 Lat. (DD):

34.47295 Long. (DD): -89.257517 T-6S, R-1W, S-13, Q-E.

Stream order: 5. Link magnitude: 108.

West Cypress Creek, Cypress Creek trib., Lafayette Co.

Date: 6/4/99. CR 251, 21.9 ENE Oxford, 12.1 km NNE

Denmark. Puskus quad. Lat. (DMS): N34-24-31.5 Long.

(DMS): W89-17-34.1 Lat. (DD): 34.40525 Long. (DD): -

89.289017 T-8S, R-1W, S-3/10, Q-SW/NW. Stream order: 3.

Link magnitude: 16.

West Cypress Creek, Cypress Creek trib., Lafayette Co.

Date: 6/1/99. FS 844C, 18.4 km ENE Oxford, 9.2 km N

Denmark. Puskus quad. Lat. (DMS): N34-23-29.2 Long.

(DMS): W89-19-36.8 Lat. (DD): 34.3882 Long. (DD): -89.3228

T-8S, R-1W, S-17, Q-NW. Stream order: 3. Link magnitude: 8.

*West Cypress Creek, Cypress Creek trib., Lafayette Co.

Date: 6/1/99. CR 252, 7.7 km N Denmark, 14.8 km E Oxford.

Puskus quad. Lat. (DMS): N34-23-04.8 Long. (DMS): W89-

20-06.7 Lat. (DD): 34.384133 Long. (DD): -89.33445 T-8S, R1W, S-18, Q-SE. Stream order: 2. Link magnitude: 12.

*Unnamed trib. West Cypress Creek, Cypress Creek trib.,

Lafayette Co. Date: 6/3/99. FS 837E, 18.4 km ENE Oxford,

10.4 km N Denmark. Puskus quad. Lat. (DMS): N34-24-29.1

Long. (DMS): W89-20-11.3 Lat. (DD): 34.40485 Long. (DD): -

89.335217 T-8S, R-1W, S-7, Q-NE. Stream order: 2. Link

magnitude: 3.

Unnamed northern trib. W. Cypress Creek, Cypress Creek

trib., Lafayette Co. Date: 6/3/99. FS 837E, 2.3 km E Keel,

10.9 km N Denmark. Puskus quad. Lat. (DMS): N34-24-27.4

Long. (DMS): W89-20-13.7 Lat. (DD): 34.404567 Long. (DD):

-89.335617 T-8S, R-1W, S-6, Q-SE. Stream order: 1. Link

magnitude: 1.

Puskus Creek, Cypress Creek trib., Lafayette Co. Date: 6/2/

99. CR 237 (FS 830), 14.4 km ENE Oxford, 9.8 km NNW

Denmark. Puskus quad. Lat. (DMS): N34-23-43.2 Long.

(DMS): W89-22-20.7 Lat. (DD): 34.390533 Long. (DD): -

89.370117 T-8S, R-2W, S-11, Q-SE. Stream order: 2. Link

magnitude: 3.

Puskus Creek, Cypress Creek trib., Lafayette Co. Date: 6/2/

99. CR 237 (FS 830), 14.4 km ENE Oxford, 9.8 km NNW

Denmark. Puskus quad. Lat. (DMS): N34-23-43.7 Long.

(DMS): W89-22-18.3 Lat. (DD): 34.390617 Long. (DD): -

89.369717 T-8S, R-2W, S-11, Q-SE. Stream order: 3. Link

magnitude: 7.

Puskus Creek, Cypress Creek trib., Lafayette Co. Date: 7/28/

99. CR 245, 17 km NE Oxford, 15 km N Denmark. Puskus

quad. Lat. (DMS): N34-26-46.7 Long. (DMS): W89-19-58.3

Lat. (DD): 34.441117 Long. (DD): -89.326383 T-7S, R-1W, S29, Q-NW. Stream order: 4. Link magnitude: 37.

Puskus Creek, Cypress Creek trib., Lafayette Co. Date: 8/3/

99. FS 832-C, 0.8 km downstream Puskus Lake, 16.2 km

NNE Oxford, 14 km SSW Cornersville. Puskus quad. Lat.

(DMS): N34-26-41.0 Long. (DMS): W89-20-14.8 Lat. (DD):

34.440167 Long. (DD): -89.3358 T-7S, R-1W, S-30, Q-SE.

Stream order: 4. Link magnitude: 34.

Bay Springs Branch, Puskus Creek trib., Lafayette Co. Date:

9/14/99. University Mississippi Field Station, 10.2 km NE

Oxford, 12 km SE Abbeville. Bagley quad. Lat. (DMS): N34-

25-34.1 Long. (DMS): W89-23-44.3 Lat. (DD): 34.42235

Long. (DD): -89.390717 T-7S, R-2W, S-34, Q-SW. Stream

order: 1. Link magnitude: 1.

Unnamed trib. Puskus Creek, Cypress Creek trib., Lafayette

Co. Date: 7/28/99. CR 832, 1 km N Puskus Lake, 16 km N

Denmark. Puskus quad. Lat. (DMS): N34-27-02.6 Long.

(DMS): W89-21-00.2 Lat. (DD): 34.450433 Long. (DD): -

89.350033 T-7S, R-1W, S-30, Q-NW. Stream order: 2. Link

magnitude: 2.

*Little Tallahatchie Canal, Marshall-Lafayette Co. Date: 8/10/

99. Riverside recreation site, 5.8 km SSW Bethlehem, 10.7

km SW Cornersville, Bethlehem quad. Lat. (DMS): N34-31-

45.6 Long. (DMS): W89-21-58.7 Lat. (DD): 34.524267 Long.

(DD): -89.359783 T-6S, R-2W, S-25, Q-SW. Stream order:

ND. Link magnitude: ND.

*Little Tallahatchie Canal, Union Co. Date: 8/9/99.

downstream Hwy 30 bridge at Etta. Etta quad. Lat. (DMS):
180

N34-28-55.2 Long. (DMS): W89-13-29.5 Lat. (DD):

34.475867 Long. (DD): -89.221583 T-7S, R-1E, S-8, Q-SW.

Stream order: ND. Link magnitude: ND.

*Little Tallahatchie R., Lafayette Co. Date: 8/12/99. Old

channel at Riverside Recreation Area, 6.5 km SSW

Bethlehem, 10.7 km SW Cornersville. Bethlehem quad. Lat.

(DMS): N34-31-11.3 Long. (DMS): W89-21-48.3 Lat. (DD):

34.51855 Long. (DD): -89.35805 T-6S, R-2W, S-36, Q-NW.

Stream order: ND. Link magnitude: ND.

*Little Tallahatchie R., Lafayette Co. Date: 9/15/99. 1.2 km

downstream Bakers Field boat ramp, 10.8 km ESE Malone,

10.8 km ENE Abbeville. Malone quad. Lat. (DMS): N34-31-

11.2 Long. (DMS): W89-22-56.2 Lat. (DD): 34.518533 Long.

(DD): -89.376033 T-6S, R-2W, S-35/34, Q-NW/NE. Stream

order: ND. Link magnitude: ND.

*Little Tallahatchie R., Lafayette Co. Date: 9/17/99. Old

channel, north ‘Hudson’s Deer Camp’, 7 m downstream trail

crossing, 8.8 km SSE Bethlehem, 4 km SSW Cornersville.

Bethlehem quad. Lat. (DMS): N34-30-20.5 Long. (DMS):

W89-16-16.7 Lat. (DD): 34.503417 Long. (DD): -89.26945 T7S, R-1W, S-2, Q-N. Stream order: ND. Link magnitude: ND.

Lee Creek, Lafayette Co. Date: 8/5/99. 40 m upstream and

downstream CR 210 bridge, 14.5 km SSW Bethlehem, 6 km

SE Abbeville. Bagley quad. Lat. (DMS): N34-28-25.2 Long.

(DMS): W89-26-46.1 Lat. (DD): 34.470867 Long. (DD): -

89.441017 T-7S, R-2W, S-18, Q-E. Stream order: 4. Link

magnitude: 34.

Lee Creek, Lafayette Co. Date: 8/5/99. 10 m upstream CR

291 bridge, 2.8 km E Abbeville, 14 km SSW Bethlehem.

Bagley quad. Lat. (DMS): N34-29-52.1 Long. (DMS): W89-

27-26.0 Lat. (DD): 34.492017 Long. (DD): -89.454333 T-7S,

R-2W, S-6, Q-SW. Stream order: 4. Link magnitude: 48.

Mitchell Creek, Union Co. Date: 8/2/99. 5 m upstream Hwy

30 bridge, 7.9 km SSW Cornersville, 6.1 km NW Enterprise.

Etta quad. Lat. (DMS): N34-28/56.9 Long. (DMS): W89-12-

09.7 Lat. (DD): 34.47615 Long. (DD): -89.201617 T-7S, R1E, S-9, Q-SE. Stream order: 4. Link magnitude: 87.

Mitchell Creek, Union Co. Date: 8/2/99. 5 m upstream CR 12

bridge, 27 km ENE Abbeville, 9.1 km SSW Myrtle. Hickory

Flat quad. Lat. (DMS): N34-31-14.1 Long. (DMS): W89-12-

09.7 Lat. (DD): 34.519017 Long. (DD): -89.201617 T-6S, R1E, S-33, Q-Center. Stream order: 3. Link magnitude: 8.

TIPPAH RIVER DRAINAGE

Big Snow Creek, Benton Co. Date: 7/13/99. 10 m upstream

CR 607 bridge, 10 km SW Ashland, 17.8 km E Holly

Springs. Holly Springs SE quad. Lat. (DMS): N34-46-22.4

Long. (DMS): W89-15-08.3 Lat. (DD): 34.7704 Long. (DD): -

89.251383 T-3S, R-1W, S-36, Q-SE. Stream order: 5. Link

magnitude: 67.

Big Snow Creek, Benton Co. Date: 7/19/99. End FS Rd. 657-

D, 20.3 km ESE Holly Springs, 13.7 km NNW Hickory Flat.

Chilli Creek quad. Lat. (DMS): N34-43-51.6 Long. (DMS):

W89-14-13.3 Lat. (DD): 34.725267 Long. (DD): -89.23555 T4S, R-1E, S-18, Q-SE. Stream order: 5. Link magnitude: 95.

Wagner Creek, Big Snow Creek trib., Benton Co. Date: 7/12/

99. 5 m downstream CR 652 bridge, 6.4 km SSW Ashland,

21.6 km ENE Holly Springs. Ashland quad. Lat. (DMS): N34-

47-26.5 Long. (DMS): W89-12-53.6 Lat. (DD): 34.78775

Long. (DD): -89.208933 T-3S, R-1E, S-29, Q-SE. Stream

order: 4. Link magnitude: 11.

Wagner Creek, Big Snow Creek trib., Benton Co. Date: 7/12/

99. 15 m upstream CR 607 bridge, 8.9 km SSW Ashland,

20.1 km ESE Holly Springs. Ashland quad. Lat. (DMS): N34-

46-05.5 Long. (DMS): W89-13-46.1 Lat. (DD): 34.767583

Long. (DD): -89.22435 T-4S, R-1E, S-6/5, Q-NE/NW. Stream

order: 4. Link magnitude: 41.

Big Spring Creek, Marshall Co. Date: 7/29/99. 40 m

upstream CR 633 bridge, 8.5 km WSW Potts Camp, 14.5 km

SSE Holly Springs. Waterford quad. Lat. (DMS): N34-37-58.2

Long. (DMS): W89-23-49.0 Lat. (DD): 34.626367 Long. (DD):

-89.3915 T-5S, R-2W, S-22, Q-NW. Stream order: 5. Link

magnitude: 206.

Big Spring Creek, Marshall Co. Date: 8/3/99. Upstream CR

694 bridge, channel runs parallel to road, 8.3 km NNE

Abbeville, 9.9 km WSW Bethlehem. Malone quad. Lat.

(DMS): N34-33-28.3 Long. (DMS): W89-25-49.6 Lat. (DD):

34.554717 Long. (DD): -89.424933 T-6S, R-W, S-17, Q-SE.

Stream order: 5. Link magnitude: 314.

Unnamed trib. Big Spring Creek, Marshall Co. Date: 7/30/99.

Musgray Rd off CR 633, upstream and downstream bridge,

9.3 km ENE Potts Camp, 11 km S Holly Springs. Waterford

quad. Lat. (DMS): N34-39-48.4 Long. (DMS): W89-24-46.1

Lat. (DD): 34.658067 Long. (DD): -89.407683 T-5S, R-2W, S9, Q-NW. Stream order: 4. Link magnitude: 71.

Chilli Creek, Benton Co. Date: 7/1/99. 10 m upstream CR

625 bridge, 15 km SSW Ashland, 9.7 km NNW Hickory Flat.

Chilli Creek quad. Lat. (DMS): N34-42-10.2 Long. (DMS):

W89-12-08.3 Lat. (DD): 34.7017 Long. (DD): -89.201383 T4S, R-1E, S-28, Q-S. Stream order: 5. Link magnitude: 91.

Chewalla Creek, Marshall Co. Date: 6/16/99. 25 m upstream

Hwy 4 bridge, 5 km S Hudsonville, 8.9 km NE Holly Springs.

Holly Springs SE quad. Lat. (DMS): N34-41-46.8 Long.

(DMS): W89-22-03.4 Lat. (DD): 34.691133 Long. (DD): -

89.367233 T-3S, R-2W, S-24/13, Q-NW/SW. Stream order: 3.

Link magnitude: 10.

Chewalla Creek, Marshall Co. Date: 6/17/99. 15 m upstream

bridge Lacy Ivy Rd. bridge, 3.9 km SE Lake Center, 4.5 km

NNW Potts Camp. Potts Camp quad. Lat. (DMS): N34-40-

44.0 Long. (DMS): W89-19-56.3 Lat. (DD): 34.674 Long.

(DD): -89.32605 T-5S, R-1W, S-5, Q-NW. Stream order: 5.

Link magnitude: 189.

Chewalla Creek, Marshall Co. Date: 6/21/99. FS Rd. 661-A

end, upstream and downstream, 3.8 km W Lake Center, 6.8

km NNW Potts Camp. Potts Camp quad. Lat. (DMS): N34-

41-52.5 Long. (DMS): W89-19-49.0 Lat. (DD): 34.692083

Long. (DD): -89.324833 T-4S, R-1W, S-29/32, Q-SW/NW.

Stream order: 5. Link magnitude: 150.

Chewalla Creek, Marshall Co. Date: 6/14/99. 100 m

downstream Chewalla Lake outflow, 8.8 km NNW Potts
181

Camp, 10.6 km SE Holly Springs. Potts Camp quad. Lat.

(DMS): N34-43-23.3 Long. (DMS): W89-20-27.8 Lat. (DD):

34.72055 Long. (DD): -89.337967 T-4S, R-1W, S-19, Q-NE.

Stream order: 4. Link magnitude: 90.

Unnamed trib. Chewalla Creek, Marshall Co. Date: 7/8/99. 10

m upstream CR 611, 13.9 km ESE Holly Springs, 16.8 km

NW Hickory Flat. Potts Camp quad. Lat. (DMS): N34-43-58.8

Long. (DMS): W89-18-12.1 Lat. (DD): 34.726467 Long. (DD):

-89.302017 T-4S, R-1W, S-16, Q-SE. Stream order: 3. Link

magnitude: 11.

Chilli Creek, Benton Co. Date: 6/30/99. 30 m downstream

North Chilli Lake outflow. Chilli Creek quad. Lat. (DMS): N34-

42-00.5 Long. (DMS): W89-08-23.2 Lat. (DD): 34.700083

Long. (DD): -89.1372 T-4S, R-2E, S-30, Q-SW. Stream order:

3. Link magnitude: 8.

East Chilli Creek, Chilli Creek trib., Benton Co. Date: 6/30/

99. CR 626E downstream Chilli Lake, 40 m upstream

confluence southern trib. Chilli Creek quad. Lat. (DMS): N34-

41-46.0 Long. (DMS): W89-08-52.4 Lat. (DD): 34.691 Long.

(DD): -89.142067 T-4S, R-1E, S-36, Q-NE. Stream order: 3.

Link magnitude: 12.

Curtis Creek, Benton Co. Date: 7/22/99. FS Rd. 609A, 30 m

downstream Curtis Lake outflow, 8.5 km SE Ashland, 7 km S

Yellow Rabbit Lake. Whitten Town quad. Lat. (DMS): N34-46-

44.5 Long. (DMS): W89-06-38.1 Lat. (DD): 34.774083 Long.

(DD): -89.10635 T-3S, R-2E, S-32, Q-NE. Stream order: 3.

Link magnitude: 2.

Oaklimeter Creek, Benton Co. Date: 6/29/99. 40 m

downstream Wood Duck Lake outflow, 7.8 km SSE Bethel,

18 km E Potts Camp. Blue Mountain quad. Lat. (DMS): N34-

39-59.0 Long. (DMS): W89-06-19.4 Lat. (DD): 34.659833

Long. (DD): -89.103233 T-5S, R-2E, S-8/9, Q-NE/NW.

Stream order: 3. Link magnitude: 12.

Oaklimeter Creek, Benton Co. Date: 6/29/99. 15 m upstream

CR 626 bridge, 7.1 km SSE Bethel, 18 km E Potts Camp.

Blue Mountain quad. Lat. (DMS): N34-39-39.8 Long. (DMS):

W89-06-28.1 Lat. (DD): 34.656633 Long. (DD): -89.104683

T-5S, R-2E, S-8, Q-SE. Stream order: 4. Link magnitude: 21.

Oaklimeter Creek, Benton Co. Date: 7/6/99. 15 m upstream

CR 640 bridge, 8.1 km WNW Hickory Flat, 3.7 km SE Potts

Camp. Potts Camp quad. Lat. (DMS): N34-37-42.5 Long.

(DMS): W89-16-27.3 Lat. (DD): 34.62375 Long. (DD): -

89.271217 T-5S, R-1W, S-23, Q-S. Stream order: 5. Link

magnitude: 277.

Pechahallee Creek, Oaklimeter Creek trib., Benton Co. Date:

6/23/99. 12 m upstream FS 692 bridge, 5 km W Hickory Flat,

6.5 km SE Potts Camp. Hickory Flat. quad. Lat. (DMS): N34-

37-12.1 Long. (DMS): W89-14-24.8 Lat. (DD): 34.618683

Long. (DD): -89.237467 T-5S, R-1E, S-30, Q-NW. Stream

order: 4. Link magnitude: 31.

*Pechahallee Creek, Oaklimeter Creek trib., Benton Co.

Date: 6/25/99. 30 m downstream Hwy 5 bridge, 4.8 km N

Hickory Flat, 8 km SSE Bethel. Chilli Creek quad. Lat.

(DMS): N34-39-39.1 Long. (DMS): W89-11-03.5 Lat. (DD):

34.656517 Long. (DD): -89.183917 T-5S, R-1E, S-10, Q-S.

Stream order: 1. Link magnitude: 1.

Pechahallee Creek, Oaklimeter Creek trib., Benton Co. Date:

7/2/99. 15 m upstream CR 625 bridge, 4 km NW Hickory

Flat, 8.3 km E Potts Camp. Chilli Creek quad. Lat. (DMS):

N34-38-21.2 Long. (DMS): W89-13-05.3 Lat. (DD):

34.636867 Long. (DD): -89.21755 T-5S, R-1E, S-20, Q-NE.

Stream order: 4. Link magnitude: 21.

Unnamed trib. Oaklimeter Creek, Benton Co. Date: 6/25/99.

10 m upstream box culvert opening on north side Hwy 78,

9.3 km ESE Potts Camp, 2.5 km NNW Hickory Flat. Chilli

Creek quad. Lat. (DMS): N34-37-43.9 Long. (DMS): W89-12-

29.4 Lat. (DD): 34.623983 Long. (DD): -89.2049 T-5S, R-1E,

S-21, Q-SW. Stream order: 2. Link magnitude: 3.

*Unnamed trib. Oaklimeter Creek, Benton Co. Date: 7/7/99.

75 m upstream CR 662, 4.3 km E Potts Camp, 7.4 km WNW

Hickory Flat. Potts Camp quad. Lat. (DMS): N34-38-20.9

Long. (DMS): W89-15-45.6 Lat. (DD): 34.636817 Long. (DD):

-89.2576 T-5S, R-1W, S-24, Q-NW. Stream order: 1. Link

magnitude: 1.

*Potts Creek, Benton Co. Date: 6/22/99. 3 m downstream CR

638, downstream Brents Lake, 6.4 km NNW Cornersville, 7

km SSE Potts Camp. Bethlehem quad. Lat. (DMS): N34-35-

13.2 Long. (DMS): W89-16-46.1 Lat. (DD): 34.585533 Long.

(DD): -89.27435 T-6S, R-1W, S-2, Q-SW. Stream order: 4.

Link magnitude: 40.

Potts Creek, Marshall Co. Date: 6/23/99. 15 m downstream

Mills Rd. bridge, 2.5 km NNW Bethlehem, 7 km SSW Potts

Camp. Bethlehem quad. Lat. (DMS): N34-35-30.6 Long.

(DMS): W89-20-32.5 Lat. (DD): 34.588433 Long. (DD): -

89.33875 T-6S, R-1W, S-6, Q-NW. Stream order: 5. Link

magnitude: 142.

Unnamed trib. Potts Creek, Benton Co. Date: 6/22/99. 7 m

upstream 347 bridge, 4.4 km NE Bethlehem, 10 km W

Hickory Flat. Bethlehem quad. Lat. (DMS): N34-36-13.2

Long. (DMS): W89-18-55.6 Lat. (DD): 34.6022 Long. (DD): -

89.309267 T-5S, R-1W, S-33, Q-SW. Stream order: 4. Link

magnitude: 21.

Unnamed trib. Tippah R., Benton Co. Date: 6/17/99. 20 m

downstream CR 657 bridge, 7.2 km NNE Potts Camp, 10.8

km W Bethel. Potts Camp quad. Lat. (DMS): N34-42-32.7

Long. (DMS): W89-15-16.5 Lat. (DD): 34.70545 Long. (DD): -

89.25275 T-4S, R-1W, S-25, Q-NE. Stream order: 4. Link

magnitude: 26.

Unnamed trib. Tippah R., Benton Co. Date: 6/17/99. 110 m

upstream CR 606, Gandy Property, 4.4 km NNE Potts

Camp, 7.3 km E Pine Grove. Potts Camp quad. Lat. (DMS):

N34-40-50.6 Long. (DMS): W89-16-48.3 Lat. (DD): 34.6751

Long. (DD): -89.274717 T-5S, R-1W, S-2, Q-NE. Stream

order: 4. Link magnitude: 36.

*Unnamed trib. Tippah R., Benton Co. Date: 7/9/99. 8 m

upstream trail crossing and upstream confluence,

unnumbered FS rd., 1.7 km NE Potts Camp. Potts Camp

quad. Lat. (DMS): N34-39-35.4 Long. (DMS): W89-17-08.3
182

Lat. (DD): 34.6559 Long. (DD): -89.284717 T-5S, R-1W, S10, Q-SE. Stream order: 3. Link magnitude: 4.

Unnamed trib. Tippah R., Benton Co. Date: 7/9/99. 5 m

downstream confluence, 1.7 km NE Potts Camp. Potts Camp

quad. Lat. (DMS): N34-39-36.5 Long. (DMS): W89-17-10.9

Lat. (DD): 34.656083 Long. (DD): -89.28515 T-5S, R-1W, S10, Q-SE. Stream order: 3. Link magnitude: 7.

Unnamed trib. Tippah R., Benton Co. Date: 7/7/99. FS Rd.

606-L, 30 m downstream Cox Lake outflow, 5.3 km NE Potts

Camp, 17.1 km SE Holly Springs. Potts Camp quad. Lat.

(DMS): N34-41-28.3 Long. (DMS): W89-17-24.5 Lat. (DD):

34.68805 Long. (DD): -89.287417 T-4S, R-1W, S-34, Q-S.

Stream order: 3. Link magnitude: 25.

Yellow Rabbit Creek, Benton Co. Date: 7/21/99. 3 m

downstream CR 649 bridge, 6.3 km ESE Ashland, 16.4 km

WNW Ripley. Whitten Town quad. Lat. (DMS): N34-49.09.3

Long. (DMS): W89-06-19.0 Lat. (DD): 34.818217 Long. (DD):

-89.103167 T-3S, R-2E, S-17/16, Q-SE/SW. Stream order: 3.

Link magnitude: 8.

Yellow Rabbit Creek, Benton Co. Date: 7/22/99. 80 m

upstream CR 648 bridge, 7 km SSE Ashland, 15 km WNW

Ripley. Ashland quad. Lat. (DMS): N34-46-25.8 Long. (DMS):

W89-08-41.7 Lat. (DD): 34.770967 Long. (DD): -89.140283 T3S, R-1E, S-36, Q-SE. Stream order: 4. Link magnitude: 38.

WOLF RIVER DRAINAGE

Indian Creek, Benton Co. Date: 7/14/99. 5 m downstream

CR 646 (Blackjack Rd.), 9 km NE Sahland, 5.5 km NNE

Yellow Rabbit Lake. Whitten Town quad. Lat. (DMS): N34-51-

46.2 Long. (DMS): W89-05-08.0 Lat. (DD): 34.8577 Long.

(DD): -89.084667 T-2S, R-2E, S-34, Q-SW. Stream order: 3.

Link magnitude: 9.

Sourwood Creek, Benton Co. Date: 7/15/99. 10 m upstream

CR 647 bridge, 9.3 km SSE Canaan, 17.8 km NW Benton.

Camp Hill quad. Lat. (DMS): N34-52-39.4 Long. (DMS): W89-

04-11.3 Lat. (DD): 34.873233 Long. (DD): -89.06855 T-2S, R2E, S-27/26, Q-SE/SW. Stream order: 3. Link magnitude: 7.

Sourwood Creek, Benton Co. Date: 7/15/99. 8 m

downstream CR 646 bridge, 7.3 km SE Canaan, 19.3 km

NNW Ripley. Camp Hill quad. Lat. (DMS): N34-53-17.1 Long.

(DMS): W89-04-49.4 Lat. (DD): 34.886183 Long. (DD): -

89.0749 T-2S, R-2E, S-22/27, Q-NE. Stream order: 3. Link

magnitude: 17.

Turkey Creek, Benton Co. Date: 8/11/99. Blackburn Rd off

CR 647, 2 m upstream bridge, 12.5 km NNE Ashland, 19.2

km NNW Ripley. Camp Hill quad. Lat. (DMS): N34-53-47.0

Long. (DMS): W89-03-36.6 Lat. (DD): 34.891167 Long. (DD):

-89.0561 T-2S, R-2E, S-23, Q-NE. Stream order: 3. Link

magnitude: 30.

YOCONA RIVER DRAINAGE

Kettle Creek, Lafayette Co. Date: 6/10/99. 80 m downstream

Drewery Lake outflow, 17.3 km E Oxford, 4 km N Denmark.

Denmark quad. Lat. (DMS): N34-20-41.8 Long. (DMS): W89-

20-12.3 Lat. (DD): 34.3403 Long. (DD): -89.335383 T-8S, R1W, S-31, Q-NE. Stream order: 2. Link magnitude: 2.

Kettle Creek, Lafayette Co. Date: 6/10/99. 120 m upstream

CR 277 bridge, 18.1 km ESE Oxford, 4 km ENE Denmark.

Denmark quad. Lat. (DMS): N34-20-07.1 Long. (DMS): W89-

19-53.5 Lat. (DD): 34.334517 Long. (DD): -89.325583 T-8S,

R-1W, S-32, Q-SW. Stream order: 3. Link magnitude: 4.

Kettle Creek, Lafayette Co. Date: 6/8/99. 10 m upstream Hwy

6 bridge, 18.4 km ESE Oxford, 1.7 km ENE Denmark.

Denmark quad. Lat. (DMS): N34-18-48.6 Long. (DMS): W89-

19-42.1 Lat. (DD): 34.3081 Long. (DD): -89.323683 T-9S, R1W, S-8, Q-NW. Stream order: 3. Link magnitude: 9.

Pumpkin Creek, Lafayette Co. Date: 6/11/99. 10 m upstream

Hwy 6 bridge, 12 km ESE Oxford, 5.7 km NNE Yocona.

Yocona quad. Lat. (DMS): N34-19-37.5 Long. (DMS): W89-

23-51.3 Lat. (DD): 34.322917 Long. (DD): -89.391883 T-9S,

R-2W, S-3, Q-NW. Stream order: 3. Link magnitude: 11.

Pumpkin Creek, Lafayette Co. Date: 6/11/99. Past end CR

266, 2.2 km upstream Hwy 6 bridge, 7.4 km ESE Oxford, 7.4

km N Yocona. Yocona quad. Lat. (DMS): N34-20-20.8 Long.

(DMS): W89-23-02.2 Lat. (DD): 34.3368 Long. (DD): -

89.3837 T-8S, R-2W, S-35, Q-NW. Stream order: 3. Link

magnitude: 6.

Yellow Leaf Creek, Lafayette Co. Date: 6/8/99. 40 m

upstream CR 225, 6.7 km ESE Oxford, 10.4 km NNW

Yocona. Yocona quad. Lat. (DMS): N34-22-04.4 Long. (DMS):

W89-25-44.1 Lat. (DD): 34.3674 Long. (DD): -89.424017 T8S, R-2W, S-20, Q-SE. Stream order: 3. Link magnitude: 9.

Yellow Leaf Creek, Lafayette Co. Date: 6/9/99. Deloach

property, 10 m upstream from trib. confluence, 8.2 km E

Oxford, 16.3 km SE Abbeville. Bagley quad. Lat. (DMS):

N34-22-32 Long. (DMS): W89-25-17 Lat. (DD): 34.372 Long.

(DD): -89.4195 T-8S, R-2W, S-20/21, Q-NE/NW. Stream

order: 2. Link magnitude: 2.

Yellow Leaf Creek, Lafayette Co. Date: 6/9/99. Deloach

property, 100 m downstream from trib confluence, 8.2 km E

Oxford, 16.2 km SE Abbeville. Yocona quad. Lat. (DMS):

N34-22-27 Long. (DMS): W89-25-17 Lat. (DD): 34.371167

Long. (DD): -89.4195 T-8S, R-2W, S-20/21, Q-NE/NW.

Stream order: 3. Link magnitude: 8.

Yellow Leaf Creek, Lafayette Co. Date: 6/9/99. Deloach

property, 1.2 km upstream from trib. confluence, 8.7 km E

Oxford, 16.3 km SE Abbeville. Bagley quad. Lat. (DMS):

N34-22-48 Long. (DMS): W89-24-38 Lat. (DD): 34.374667

Long. (DD): -89.406333 T-8S, R-2W, S-16, Q-SE. Stream

order: 3. Link magnitude: 6.

Yellow Leaf Creek, Lafayette Co. Date: 7/21/99. FS 849E, 5

m downstream confluence, 17.8 km SW Etta, 17 km SE

Abbeville. Bagley quad. Lat. (DMS): N34-23-01.1 Long.

(DMS): W89-24-06.9 Lat. (DD): 34.383517 Long. (DD): -

89.40115 T-8S, R-2W, S-15, Q-W. Stream order: 2. Link

magnitude: 2.
183

INTRODUCTION

The land-water interface region of aquatic ecosystems is

always the most productive per-unit area along the gradient

from land to open water. Because most aquatic ecosystems

occur in terrain of gentle slopes and are small (mean area <

100 ha) and shallow (mean < 5 m), the wetland-littoral

components usually dominate in productivity and the

synthesis of organic matter (Wetzel 1990, 1992). Because of

the predominance of small, shallow freshwater bodies, most

dissolved organic carbon (DOC) of lacustrine and riverine

ecosystems is derived from photosynthesis of higher plants

and microflora associated with detritus, including sediments,

and is only augmented by releases of organic matter

produced by phytoplankton.

The dissolved organic compounds generated by the

production and partial degradation of organic matter

produced in terrestrial soils, wetland, and littoral interface

regions move downgradient toward the open-water regions

of the lake or river. En route, partial utilization by largely

attached microorganisms effects a selective decomposition

of more labile organic compounds and an increase in

organic substrate recalcitrance.

Dissolved and colloidal organic matter of inland waters can

be separated on the basis of many criteria. Although

separated from particulate organic matter by filtration at a

size of 0.5 mm, for an array of practical reasons (Wetzel and

Likens 2000), dissolved organic matter is often further

separated by molecular-sized categories (Cabaniss and

others 2000, Gustafsson and Gschwend 1997). The

molecular weight of humic substances influences their

proton and metal-binding characteristics, partitioning

behavior, and adsorption properties. Average molecular

weight distributions of aquatic fulvic acids vary from 2.7 to

3.0 kD and include low-MW fractions of 1,110 D and a highMW fraction of ca. 4,890 D. Less that 10 percent of dissolved

humic substances were found to be greater than 5 kD. The

low-MW fractions are more hydrophilic, exhibit faster

diffusion and increased mobility, and are commonly more

bioavailable. In contrast, the high-MW components have

increased uptake of hydrophobic organic compounds,

greater aromatic components, decreased mobility and

diffusion, and are usually appreciably less susceptible to

biological utilization. Molecular-sized fractionation helps

little, however, in relation to bioavailability of different

compounds, particularly above the size of effective cellular

membrane permeability of ca. 500 D. Bonding structure,

particularly of aromatic components, and availability of sites

for enzyme hydrolysis are of greater importance. The

present study addresses aspects that alter bioavailability of

these humic compounds as they are imported to lakes and

rivers.

Because of the relatively high utilization and turnover rates

of simple, low molecular weight compounds by microbes, as

much as 80 percent of dissolved organic matter in inland

waters is composed of organic acids. Many of these

compounds originate from higher aquatic and terrestrial

plants. Of these organic acids, some 30 to 40 percent are

composed of aromatic carbon compounds originating from

structural plant tissues (Malcolm 1990).

Because of their large size and limited accessibility of large

portions of these molecules to enzymatic hydrolysis,

degradation rates of many of these commonly acidic

macromolecules is slow. As a result, these heterogeneous

compounds exhibit long turnover times with relatively long

environmental residence times. Dissolved macromolecules

are often of considerable age (months) but are mixed with

variable and rapidly changing inputs of younger humic and

nonhumic substances. Recent studies indicated that humic

substances, particularly relatively recalcitrant fulvic acids,

are also generated by algae and decomposing

ORIGINS, FATES, AND RAMIFICATIONS OF NATURAL ORGANIC

COMPOUNDS OF WETLANDS

Robert G. Wetzel1

1 Bishop Professor of Biological Sciences, University of Alabama, Tuscaloosa, AL 35487-0206, Present address: Professor, Environmental Sciences and Engineering, The University of North Carolina, Chapel Hill, NC 27599-7431.

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability of

wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.

Abstract—Much of the organic carbon for heterotrophic metabolism in aquatic ecosystems is soluble and derived from

structural compounds of higher plants of terrestrial and wetland-littoral sources of both lake and river ecosystems. The

chemical recalcitrance of this organic matter and its oxidative utilization are fundamentally different from many sources

within the aquatic ecosystems. Within the lake or river, complex physical interactions occur that can greatly modify rates of

utilization and biochemical reactions. Natural photolysis by photosynthetically active radiation, UV-A and UV-B, can result in

the partial degradation of these macromolecules, reactivate complexed enzymes, and generate simple organic compounds

and nutrients. A portion of the dissolved organic matter is photolytically degraded completely to CO2. The chemical

recalcitrance of these organic compounds (a) represents a fundamental, often dominant, subsidy of organic matter that

drives metabolism in fresh waters and (b) is an essential aspect of metabolic stability in aquatic ecosystems. Climatic

changes affect this metabolic stability in both positive and negative ways but, generally, will increase instabilities.
184

microorganisms, and these contribute to the multitude of

diverse compounds of the composite dissolved organic

matter. This latter source is particularly important in wetland

and littoral areas, where attached algal productivity is often

many orders of magnitude greater than that of

phytoplankton, e.g., review of Wetzel 1996.

Early detailed studies of fluxes of organic carbon in lake

ecosystems indicated both that dissolved organic matter did

not accumulate or precipitate, and that large quantities of

CO2

 evaded to the atmosphere, e.g., Cole and others 1994,

Kling and others 1992, Otsuki and Wetzel 1974, Wetzel

1995). The CO2

 evasion of respiratory origins within the

lakes was always greatly in excess of autochthonous

photosynthetic organic carbon production by phytoplankton.

In spite of this evidence, for decades in aquatic ecology, the

apparent chemical recalcitrance of fulvic and humic

hydrophobic organic substances that dominated the

instantaneous bulk DOC of standing and running waters was

believed to result in poor utilization by microbiota. Loss rates

are slow but consistently in the range of 0.5 to 2 percent per

day under many different environmental conditions and,

often, are faster.

A number of studies have demonstrated that ultraviolet

irradiance (UV) can photolyse portions of proteinaceous and

humic macromolecules, e.g., Moran 2000, reviews of Moran

and Zepp 1997, Wetzel and others 1995. The resulting

dissolved organic substances generated by photolysis

resulted in immediate stimulation of and sustained bacterial

growth, e.g., Lindall and others 1995; Moran and Zepp 1997;

Stewart and Wetzel 1981, 1982; Wetzel and others 1995.

Chemical analyses of these phytolytic transformations

indicated that small organic fractions, particularly numerous

small fatty acids that serve as excellent bacterial substrates,

were generated by partial photolysis of the humic

substances.

In the present study, dissolved organic matter leached from

decomposing foliage of several wetland and floodplain

plants was examined at intervals during their individual

decomposition over an annual period. The dissolved organic

matter was exposed to different durations of spectral

fractions of photosynthetically active radiation, UV-A, and

UV-B of natural sunlight. Photolytic products of organic

matter and CO2

 were examined in relation to bacterial

growth on these products under standardized growth

conditions with natural wetland bacterial consortia.

MATERIALS AND METHODS

Dissolved organic compounds were isolated from whole

leachates from decomposing foliage of emergent littoral

aquatic plants and floodplain trees. These plants are

dominant species in the Talladega Wetland Ecosystem

(TWE) of the Talladega National Forest, Hale County, AL

(Mann and Wetzel 1995, 2000). Emergent aquatic plants

included: (a) the common cattail (Typha latifolia L.), which

exhibits four or more cohorts per year (Dickerman and

Wetzel 1985) and much standing dead tissue that

undergoes considerable leaching of DOC; and, (b) the spike

rush (Juncus effusus L.), a dominant within the TWE that

exhibits at least seven cohorts per year at this latitude and

contains much standing dead tissue that is constantly

leaching DOC (Mann and Wetzel 1996, Wetzel and Howe

1999). Floating leaves of the common water lily (Nymphaea

odorata Aiton) served as a representative of a rapidly

decaying species that exhibits a leaf turnover rate of about

every 30 days throughout the active growing season (March

through November) at this latitude [Carter, S.; Ward, G.M.;

Wetzel, R.G.; Benke, A.C. Growth, production, and

senescence of Nymphaea odorata Aiton in a southeastern

(U.S.A.) wetland. Manuscript in review. Aquatic Botany].

Leaves of the common semiaquatic alder [Alnus serrulata

(Ait.) Willd.] were also collected during late summer and

autumnal abscission. Alder is well known for rapid leaching

of DOC and fast rates of decomposition in streams. Leaves

of the riverine red oak (Quercus rubra L.) are retained for

long periods after senescence, and usually abscise in spring

(February at this latitude) after lengthy periods of slow

leaching of DOC during the autumnal and winter rainy

period.

Combined fresh and partly senesced culms and leaves of

the two emergent and the floating-leaved aquatic

angiosperms and mature leaves of the semiaquatic alder

and riparian oak were separately decomposed anaerobically

in closed chambers at 20 °C in the laboratory. At different

intervals over a period of decomposition of over 2 years,

prefiltered (precombusted at 500 °C Whatman GF/F glass

filters, 0.6-μm pore size) whole leachate was then sterile

filtered (0.2-μm pore size) and held aseptically at 4 °C until

use. Final concentrations were in the range of 10 to 20 mg

DOC-C l-1 to simulate natural concentrations, as determined

by a Shimadzu TOC-5000 analyzer.

DOC solutions were transferred aseptically to gas-tight

sealed, sterilized UV-transparent quartz chambers (4 cm in

diameter by 50 cm in length, curved upward at one end) that

contained ports for the periodic sampling of water and of

gases in the headspace above the water. The lower onethird of the tubes was incubated in a water bath to maintain

temperatures constant at 25 °C but not to interfere with light

penetration into the solutions. Four conditions were assayed:

(a) full sunlight (photosynthetically-active radiation, 720 to

400 nm (PAR), + UV-A at 400 to 320 nm + UV-B at 320 to

285 nm; (b) PAR + UV-A only; (c) PAR only; and (d) darkness

(Al foil double wrapped). Filters used for these separations

were Pyrex glass, Mylar D, and Plexiglas (Acrylite OP-2) (fig.

1). Quartz glass absorbed essentially none of the natural UV

and PAR light above 240 nm. Samples for organic

compounds in the water and CO2

 in the overlying headspace

were removed at 30-minute intervals throughout the

incubation periods of 6 hours (usually 09.00 to 15.00 hours).

Organic acids were analyzed by high performance liquid

chromatography (HPLC) collected through Teflon

microtubing from the quartz incubation chambers at time

zero and each subsequent interval. HPLC methodology

follows that described in Wetzel and others (1995). Of the

numerous fatty acids generated by photolysis of the humic

macromolecules, calibration with known compounds of

highest available quality in identical media allowed

quantitative evaluations. All error estimates are standard

deviations (± SD) based on separate analyses or

incubations for each variable and time interval (minimally

n = 3 to 5).
185

CO2

 gas samples (250 μl) were analyzed with a Varian gas

chromatograph (series 3400) equipped with a flame

ionization detector (Hayes Sep N column) and an auxiliary

oven containing a small nickel catalyst column (column: 60

°C; injector: 150 °C; auxiliary oven: 390 °C; detector: 180 °C;

duration: 3 minutes). Calibration was against a standard gas

mixture (405 ppm CO2

).

Natural consortia of bacteria were obtained from the TWE,

inoculated into modified lakewater medium (inorganic

components only, pH 7.0, slightly modified from Veerkamp

and others 1980), and grown for 48 hours at 25 °C. Eight

hundred μl of the consortium was pipetted into sterile

cryogenic (Nunc) vials with 200 μl of sterile glycerol. Tubes

were mixed, frozen, and stored in liquid nitrogen (Sambrook

and others 1989). These parent cultures were used in all

experimentation. For experiments, one frozen tube of the

consortium was thawed and added to 150 ml of modified

lakewater medium, grown at 25 °C for 24 hours and pelleted

(10,000 rpm, 10 minutes). The supernatant was decanted

and the pellet washed with 20 mM MOPS (pH 7), a

procedure that was repeated twice with MOPS and once

with sterile ultrapure water (Millipore Q-pyrogen-free). The

final pellet was suspended in 5 ml of sterile Q-water.

Dissolved organic C (DOC) content of the Q-water was

below detection by high temperature combustion with a

Shimadzu TOC-5000 analyzer. Of the final culture

suspension, 100 μl was inoculated aseptically into each

experimental flask of UV-treated DOC water.

Bacterial production was determined on nonirradiated (time

0) and sunlight-irradiated DOM of leachates of known DOC

concentrations from the known time intervals of progressive

decomposition over a period of a year. Bacterial protein

production (biomass production) was determined with the

[

3

H]leucine uptake and conversion to protein technique with

0.5-hour incubations (dark, 20 °C) on subsamples taken

from cultures immediately (day 0) and at 24-hour intervals

thereafter for several days. Methods follow Wetzel and

Likens (2000) with very high specific activity [3

H]leucine

(range of many batches 5 to 7 TBq mmole-1 or 35 to 47 GBq

mg-1) at a final concentration of 10 nM. Calculations of

protein production follow Wetzel and Likens (2000).

RESULTS AND DISCUSSION

Photolytic alterations of the recalcitrant dissolved organic

compounds were manifested in two important ways, both of

which have major significance to the metabolism of organic

matter within the ecosystems.

Partial Photolysis of Humic Substances

to Simple Organic Compounds

Photolytic changes to DOM of whole leachates, as well as

humic and fulvic acid fractions separated by ion

chromatography, released from decomposing plant

materials, were examined by solid-state 13C nuclear

magnetic resonance and low-temperature pyrolytic gas

chromatography-mass spectrometry before and after

photolysis. These analyses revealed subtle but important

qualitative changes to the bulk DOM and small but

progressive declines in DOM quantity during photolysis

(Wetzel and others 1995). Small organic fractions generated

by natural photolysis of DOM leachates showed marked,

progressively increasing release of numerous small fatty

acids, particularly acetic, levulinic, propionic, pyruvic, formic,

and citric tartaric, among others. This common pattern

resulted from photolysis by natural sunlight or if controlled

UV-A only (Pyrex filtered) light source of leachates from

different plant species. An example of the phytolytic

generation of fatty acid substrates from humic substance

leachates of the aboveground foliage of the rush Juncus

decomposing over an annual period (fig. 2). Levulinic and

formic acids were generated consistently from humic

substances over the annual period, whereas acetic,

propionic, and pyruvic/malic acids were generated more

Figure 1—Selective UV absorption of Pyrex glass, Mylar D, and

Plexiglas (Acrylite OP–2).

Figure 2—Fatty acid substrates occurring in the solutions of DOM

(20 mg C l-1) leachate (0.6-μm pore-size filtrate) from decomposing

Juncus effusus at intervals over an annual period during exposure

(minutes) to UV irradiance of increasing durations (minutes). Pyrex

glass shielding simulating UV-A and PAR portions of the spectra.
186

strongly in earlier stages of decomposition and were less

effectively cleaved from the macromolecules after several

months of decomposition.

Changes in the photolytic generation of fatty acids from the

humic substances leached from other plant species were

significantly different from those patterns seen in Juncus

(Wetzel 2000). Fatty acids generated from photolysis of

cattail (Typha) leachate exhibited persistent generation of

acetic and formic acids over the annual period of

decomposition, but other substrates, particularly levulinic,

propionic, and pyruvic/malic were of lesser quantitative

importance (fig. 3). Levulinic and tartaric acids were

generated photolytically early in the decomposition

chronology but were not found among the photolytic

products of the more recalcitrant DOM.

Dissolved organic matter from leachates of decomposing

oak (Quercus) leaves was considerably more resistant to

photolytic generation of fatty acids than were those of other

species evaluated. Acetic and propionic acids were found

early in the sequence (fig. 3), but levulinic acid, a dominant

product from partial photolysis of other plant DOM sources,

was not detected. Citric and succinic acids were generated

weakly only from the most recalcitrant DOM after long

periods of microbial degradation.

It is important to note that although the UV portion of the

spectrum is of major significance in the partial photolysis of

DOM with the release of simple substrates from the

macromolecules, PAR is also contributing to the photolysis

and generation of simple organic compounds. For example,

one of the dominant fatty acids, levulinic acid, released from

the photolysis of leachates from Juncus is quite effectively

generated by PAR only (fig. 4). These results indicated that

over half of this substrate was generated by PAR irradiance.

Many studies have shown that bacterial production is

markedly enhanced when exposed to organic substrates

generated by solar insolation and partial photolysis of humic

substances from decomposing plant materials (Moran and

Zepp 1998, Wetzel and others 1995). Despite the

heterogeneous assemblage of simple organic substrates

generated photolytically among the many diverse plant

sources under many different stages of organic

recalcitrance, the common pattern of enhanced bacterial

productivity occurred. The photolytic effect of producing

substrates that enhanced bacterial productivity increased

with increasing age of the DOM. In the example given in

figure 5, only the bacterial production is presented in

response to photolytically-decomposed DOM of Juncus

leachates at the beginning and after 6 hours of exposure to

sunlight or controlled UV simulations. This same pattern was

found with humic macromolecules released in leachates of

the other species as well, although the rates at which they

occurred differed markedly. For example, this pattern was

found among leachate sources from the water lily

Nymphaea already at week 5, whereas among Juncus,

Quercus, and Typha the effects were most pronounced from

the more recalcitrant DOM (Wetzel 2000).

Figure 3—Fatty acid substrates occurring in the solutions of DOM

(20 mg C l-1) leachate (0.6-μm pore-size filtrate) from decomposing

Juncus effusus, Typha latifolia, Alnus serrulata, and Quercus rubra

at intervals over a 1-year period during exposure to UV-A and PAR

portions of the spectra at the maximum exposure duration (360

minutes) (from Wetzel 2000 by permission).

Figure 4—Net generation of levulinic acid from the partial photolysis

of sterile whole leachate (0.2-μm pore-size filtrate) of Juncus

effusus (10 mg C l-1) after 4 weeks of microbial decomposition at 20

°C in the dark. Exposed to full natural sunlight (total insolation over

4-hour period = 13.05 mol m-2), PAR only, or incubated simultaneously in the dark.
187

Complete Photolysis of DOM to CO2

Previous studies of the photolytic degradation of dissolved

organic matter suggested that the dominant component of

solar irradiance was UV-B and UV-A, and that

photosynthetically active radiation (PAR) above 400 nm was

of little consequence. Many of these studies, however, were

not performed under sterile conditions, and, as a result,

findings were confounded by microbial contributions to

degradation and generation of CO2. Furthermore, many of

the DOM sources of these studies had been exposed to

natural sunlight for long and noncomparable periods.

The present studies have employed recalcitrant dissolved

organic matter from decaying plant sources over long

periods (at intervals to a year or more) of many different

types. Natural sunlight was used under many different

conditions of light energy seasonally. Using the five

terrestrial and aquatic plant species as sources in nearly

200 separate experiments, the results were consistently

similar. The UV-B portion of the spectrum was always most

effective in complete photodegradation to CO2, but UV-A was

also highly effective with small differences from the

photolytic capacities of UV-B, e.g., fig. 5. Contrary to the

promulgations of others, PAR is highly effective in photolytic

degradation to CO2.

The relative effectiveness of UV-B, UV-A, and PAR in this

photodegradation to CO2 varies with species and was more

effective among leachates from relatively rapidly

decomposing plant material such as Alnus and Nymphaea.

Leachates from the cattail (Typha), rush (Juncus), and oak

(Quercus) were consistently but less effectively

photodegraded by PAR alone, e.g., fig. 6, upper. The

effectiveness of relative photolysis by UV-B, UV-A, and PAR

was more strongly correlated with species of plant DOM

origins and the relative length of time to which the DOM had

been degraded microbially before exposed to light rather

than to total light intensity received upon exposure. For

example, when natural sunlight was attenuated very rapidly,

as by a severe thunderstorm, the rate of photolytic

degradation of DOC to CO2

 declined precipitously (fig. 6,

middle), but the photolytic capacities of PAR declined more

rapidly than did the effects of UV-B (fig. 6, lower).

The high concentrations of DOM of inland waters,

predominantly humic substances of plant structural tissue

origin, serve as massive stores of relatively recalcitrant

organic carbon and energy (Wetzel 1990, 1995). Photooxidation of DOM by natural sunlight, although recognized

long ago, only recently has demonstrated the magnitude of

this process of direct utilization and alteration chemically for

enhanced microbial utilization. Findings that less energetic

spectral components, particularly those of the PAR range,

can also alter chemical availability, increase the dimensions

of these photochemical processes. PAR penetrates water to

a much greater extent than UV-A and especially UV-B and,

as a consequence, can influence a large volume of aquatic

ecosystems.

Reduction of ozone in the stratosphere and the associated

increase in UV irradiance could lead to accelerated

photolytic degradation of macromolecules of DOM by both

abiotic and biotic pathways to CO2

. In addition, the photolytic

Figure 5—Photolytic degradation of sterile Juncus whole leachate

(0.2-μm pore-size filtrate) after 4 weeks of microbial decomposition

to CO2

 under replicated, aseptic conditions exposed to full sunlight

(15.53 mol m-2 over the 4-hour period), UV-A + PAR, and PAR only. Figure 6—Photolytic degradation of sterile Juncus whole leachate

(0.2-μm pore-size filtrate) after 20 weeks of microbial decomposition

to CO2 under replicated, aseptic conditions exposed to full sunlight,

UV-A + PAR, and PAR only (upper). A severe rainstorm occurred

during the incubations, which reduced light severely for an hour

(middle). The net change in CO2 production per amount of light

received per interval under these conditions (lower).
188

enhancement of substrates for bacterial metabolism can

result in accelerated rates of biogeochemical cycling of

nutrients and stimulated productivity of the ecosystems.

Even if UV photoinhibition of microbes occurs in relatively

clear surface waters (cf. review of Karentz and others 1994),

the photolytic organic substrate products are readily

transported by water circulation to photoprotected

heterotrophic regions of greater depth. The resulting

enhanced microbial respiration will certainly lead to

enhanced generation of CO2

 and evasion to the atmosphere.

In addition, as the concentrations of CO2 in the atmosphere

increase, largely from anthropogenic combustion of fossil

fuels, plant growth commonly accelerates. Concentrations of

CO2 will double to ca. 720 ppm yet in the 21st century.

Experimental enhancement of CO2 leads to increased rates

of photosynthesis and growth by about 135 percent and

commonly leads to nitrogen limitations. As a result, C:N

ratios increase and commonly structural tissues, particularly

lignin content, increase. As these plants grown in enriched

CO2 decay, DOM released from leachates into the surface

waters contains greater amounts of dissolved humic

substances. Photolysis of this DOM shows a conspicuous

increase in CO2 release from those plants grown under

atmospherically CO2 enriched conditions among both

terrestrial plants (fig. 7) and emergent aquatic plants (fig. 8).

Clearly, photodegradation of dissolved organic substrates is

a major process altering rates of biogeochemical cycling in

aquatic ecosystems.

REFERENCES

Cabaniss, S.E.; Zhou, Q.; Maurice, P.A. [and others]. 2000. A lognormal distribution model for the molecular weight of aquatic

fulvic acids. Environmental Science and Technology. 34: 1103–

1109.

Cole, J.J.; Caraco, N.F.; Kling, G.W.; Kratz, T.K. 1994. Carbon

dioxide supersaturation in the surface waters of lakes. Science.

265: 1568–1570.

Dickerman, J.A.; Wetzel, R.G. 1985. Clonal growth in Typha latifolia:

population dynamics and demography of the ramets. Journal of

Ecology. 73: 535–552.

Gustafsson, Ö.; Gschwend, P.M. 1997. Aquatic colloids: concepts,

definitions, and current challenges. Limnology & Oceanography.

42: 519–528.

Karentz, D. [and others]. 1994. Impact of UV-B radiation on pelagic

freshwater ecosystems. Arch. Hydrobiol. Beih. Ergebn. Limnol.

36: 31–69.

Kling, G.W.; Kipphut, G.W.; Miller, M.C. 1992. The flux of CO2

 and

CH4 from lakes and rivers in arctic Alaska. Hydrobiologia. 240:

23–36.

Lindell, M.J.; Granéli, W.; Tranvik, L.J. 1995. Enhanced bacterial

growth in response to photochemical transformation of dissolved

organic matter. Limnology & Oceanography. 40: 195–199.

Malcolm, R.L. 1990. The uniqueness of humic substances in each

of soil, stream, and marine environments. Anal. Chem. Acta. 232:

Figure 7—Release of CO2 19–30. by complete photolytic degradation of

sterile leachate from leaves of the poplar tree (Populus

tremuloides), grown under ambient atmospheric CO2 and doubled

CO2

 concentrations, after 1 week of microbial decomposition.

Figure 8—Release of CO2 by complete photolytic degradation of

sterile leachate from leaves of the cattail (Typha latifolia), grown

under ambient atmospheric CO2 and doubled CO2 concentrations,

after 1 week of microbial decomposition.
189

Mann, C.J.; Wetzel, R.G. 1995. Dissolved organic carbon and its

utilization in a riverine wetland ecosystem. Biogeochemistry. 31:

91–120.

Mann, C.G.; Wetzel, R.G. 1996. Loading and bacterial utilization of

dissolved organic carbon from emergent macrophytes. Aquatic

Botany. 53: 61–72.

Mann, C.J.; Wetzel, R.G. 2000. Hydrology of an impounded lotic

wetland - subsurface hydrology. Wetlands. 20: 33–43.

Moran, M.A.; Zepp, R.G. 1997. Role of photoreactions in the

formation of biologically labile compounds from dissolved

organic matter. Limnology & Oceanography. 42: 1307–1316.

Otsuki, A.; Wetzel, R.G. 1974. Calcium and total alkalinity budgets

and calcium carbonate precipitation of a small hard-water lake.

Arch. Hydrobiol. 73: 14–30.

Sambrook, J.E.; Fritsch, F.; Maniatis, T. 1989. Storage media. In:

Molecular cloning: a laboratory manual. Pt 3. 2d ed. Maine: Cold

Spring Harbor Laboratories: A.5.

Stewart, A.J.; Wetzel, R.G. 1981. Dissolved humic materials:

photodegradation, sediment effects, and reactivity with phosphate and calcium carbonate precipitation. Arch. Hydrobiol. 92:

265–286.

Stewart, A.J.; Wetzel, R.G. 1982. Influence of dissolved humic

materials on carbon assimilation and alkaline phosphatase

activity in natural algal-bacterial assemblages. Freshwater

Biology. 12: 369–380.

Wetzel, R.G. 1990. Land-water interfaces: metabolic and limnological regulators. Verhand. Internat. Verein. Limnol. 24: 6–24.

Wetzel, R.G. 1992. Gradient-dominated ecosystems: sources and

regulatory functions of dissolved organic matter in freshwater

ecosystems. Hydrobiologia. 229: 181–198.

Wetzel, R.G. 1995. Death, detritus, and energy flow in aquatic

ecosystems. Freshwater Biology. 33: 83–89.

Wetzel, R.G. 1996. Benthic algae and nutrient cycling in standing

freshwater ecosystems. In: Stevenson, R.J.; Bothwell, M.; Lowe,

R., eds. Algal ecology: benthic algae in freshwater ecosystems.

New York: Academic Press: 641–667.

Wetzel, R.G. 2001. Natural photodegradation by UV-B of dissolved

organic matter of different decomposing plant sources to readily

degradable fatty acids. Verhand. Internat. Verein. Limnol. 27:

2036–2043.

Wetzel, R.G.; Hatcher, P.G.; Bianchi, T.S. 1995. Natural photolysis

by ultraviolet irradiance of recalcitrant dissolved organic matter

to simple substrates for rapid bacterial metabolism. Limnology &

Oceanography. 40: 1369–1380.

Wetzel, R.G.; Likens. G.E. 2000. Limnological analyses. 3rd ed. New

York: Springer-Verlag. 429 p.
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DEVELOPMENT OF A SHORT-TERM (<12 DAYS), PLANT-BASED SCREENING

METHOD TO ASSESS THE BIOAVAILABILITY, BIOCONCENTRATION, AND

PHYTOTOXICITY OF HEXAHYDRO-1,3,5- TRINITRO-1,3,5-TRIAZINE (RDX) TO

TERRESTRIAL PLANTS

Linda Winfield, Steven D’Surney, and John Rodgers1

EXECUTIVE SUMMARY

Limited amounts of information have been published on the

environmental impacts of hexahydro-1,3,5-trinitro-1,3,5-

triazine (RDX) to terrestrial plant communities. RDX is one of

the two high-explosive compounds used by the U.S. military

(Davis 1998) and classified as a priority pollutant by the U.S.

Environmental Protection Agency (USEPA). Millions of acres

of land on military installations, as well as manufacturing,

storage, and disposal sites, have been contaminated with

RDX (Jenkins 1989). Therefore, environmental risk

assessments (ERAs) are conducted to determine the

potential environmental impacts of RDX on receptors.

Research on the environmental impacts of RDX on

terrestrial plants is needed to facilitate filling data gaps and

decrease the level of uncertainty and costs associated with

ERAs on RDX.

Studies were designed and conducted to evaluate the

responses of 15 terrestrial plans to short-term (< 12 days)

and long-term (2, 4, and 6 weeks) exposures to military

grade RDX. Emphasis was placed on cover plants

(nontraditional crop plants) because they are more likely to

occur in the remote areas near military installations with

RDX-contaminated soils. The target plants included species

recommended for use in assessing toxic compounds by the

USEPA and the Organization for Economic and Cooperative

Development (Fletcher and others 1988). Additionally, the

target plants naturally occur in the regions around the

military installations, are important economically, and are a

source of food for wildlife or livestock. The objectives of the

short-term (< 12 days) studies were to: (1) identify RDX

sensitive plants for use in developing a plant-based, shortterm, screening method for assessing RDX bioavailability;

and (2) provide information useful for filling data gaps on the

environmental impacts of RDX in terrestrial plants. The

objectives of the long-term (2, 4, and 6 weeks) studies were

to develop RDX uptake coefficients and evaluate the efficacy

of experiments in the short-term study for estimating plant

responses to long-term RDX exposures. A toxicological

approach evaluating a spectrum of plant responses to RDX

exposure was used during both studies.

During the short-term studies, 10 cover plants (sanfroin,

sunflower, bush grass, Delar small Burnett, red clover, white

clover, tall fescue grass, Kentucky blue grass, orchard grass,

and rye grass) and 5 crop plants (cucumber, lettuce, rape,

corn, and wheat) were exposed to Grenada, Memphis, and

Bowdre soils amended with environmentally relevant

concentrations of RDX (0 to 4000 ppm). The assessment

parameters were percent emergence, root and shoot

lengths, and adverse developmental effects. The

experiments were concluded 2 days after 50 percent of the

control seedlings had emerged, typically < 12 days.

Statistically significant differences were measured in root

length, shoot length, or percent emergence in over 50

percent of the plants (sunflower, sanfroin, Delar small

Burnett, rape, lettuce, corn, tall fescue, and orchard grass).

However, there were no consistent patterns. The most

consistent indicators of detrimental impacts following RDX

exposures were the observed adverse developmental

effects, regardless of soil type. The observed adverse

developmental effects included: underdeveloped roots, fused

or bifurcated leaves, atypical bilateral symmetry, atypical

pigmentation, atypical seedling emergence, curled or

irregular leaf margins, necrotic lesions, chlorosis, and yellow

spots. Sunflower (a forb), sanfroin (a legume), and corn (a

cereal grass) were identified as RDX-sensitive plants. Cover

and dicot plants were generally more RDX sensitive than

monocot and crop plants. Overall, sunflower was the most

RDX sensitive of the test plants and was selected for use

during the long-term studies.

Inherent plant characteristics, seed size, and root type were

more important than taxonomic status in the development of

adverse developmental effects during the short-term studies.

The seeds of sunflower, sanfroin, and corn (47.8, 22.3, and

300.5 mg per seed, respectively) were larger than those of

the other test plants (0.5 to 6.0 mg per seed for cover plants

and 1.4 to 37.3 mg per seed for crop plants). Sunflower,

sanfroin, and corn seeds had to imbibe larger volumes of

the soil solution containing soluble (bioavailable) RDX in

order to germinate as compared to the other test plants. The

nitro groups of the RDX molecule are electron deficient

(Freeman and others 1976) whereas the cellular

biomolecules, i.e., nucleic acids and enzymes, are electron

dense. Electron-deficient molecules can participate in

electophillic substitution reactions with electron-dense

1 Winfield, Research Biologist, U.S. Engineer Research and Development Center, 3909 Halls Ferry Road, Vicksburg, MS 39180; D’Surney, Genetics

Professor, University of Mississippi, University, MS 38677; Rodgers, Director, The Institute of Wildlife and Environmental Toxicology, Clemson

University, Pendleton, SC 29670, respectively.

Citation for proceedings: Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a conference on sustainability of

wetlands and water resources: how well can riverine wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS-50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station. 191 p.
191

molecules (Gregus and Klaassen 1996). Sunflower,

sanfroin, and corn developed branched roots within 8 days

following initial RDX exposure, which increased the root

surface available for absorption of bioavailable RDX. These

factors enhanced the possibilities for interaction between

RDX and cellular biomolecules.

The efficacy of the short-term screening experiment for

estimating plant responses to long-term RDX exposure was

evaluated during the long-term (2, 4, and 6 weeks) studies.

The effects of life stage (embryos and 2-week-old seedlings)

and exposure duration were also evaluated. The adverse

developmental effects observed following short-term RDX

exposures were generally less severe when the plants were

exposed to RDX-amended Grenada soil. Therefore,

Grenada soil was used during the long-term studies to

develop RDX uptake coefficients that would be conservative.

The RDX concentrations (0, 5.8, 50, and 100 ppm) used

were low to enhance the chance for plant survival, yet

predetermined to cause adverse developmental effects. The

assessment parameters included: adverse developmental

effects, growth effects (above- and belowground biomass,

maximum shoot and root lengths, maximum stem diameter,

number of leaves, root biovolume), and bioconcentrated

RDX in leaves and flowers.

The efficacy of the short-term screening experiments for

estimating plant responses to long-term RDX exposure was

validated. The adverse developmental effects observed

during the short-term and long-term exposures were similar,

regardless of life stage or exposure duration, and included:

underdeveloped roots, fused or bifurcated leaves, atypical

bilateral symmetry, curled or irregular leaf margins, necrotic

lesions, chlorosis, and yellow spots. Additional adverse

developmental effects observed following long-term RDX

exposure were wrinkled leaf surfaces and poor seedling

development. Statistically significant differences were

measured in several of the growth parameters (biomass,

number of leaves, shoot length, root length, and root

biovolume), but there were no consistent patterns.

Statistically significant differences were measured in the

bioconcentrated RDX content of leaves using both life

stages during all exposure durations. The amount of flowers

produced was not sufficient to provide data for statistical

analysis. The highest coefficient of determination (R-square)

was calculated using data for embryos following the 6- week

exposure (0.80) and following the 4-week exposure using

the 2-week-old seedling (0.73).

The short-term screening experiments can be used for site

characterizations, risk assessments, and to evaluate

remediation activities. Additionally, they are cost effective,

simple, and require significantly less time than more

traditional approaches. The growth parameters commonly

used to evaluate herbicides, inorganic, or organic

contaminants did not adequately assess the potential for

adverse environmental impacts from RDX exposure. The

toxicological approach used during these studies was more

appropriate.

REFERENCES

Davis, W.C. 1998. Introduction to explosives. In: Jukas, J.A.; Walters,

W.P., eds. Explosive effects and applications. New York, Berlin,

Heidelberg: Springer-Verlag Publishers.

Fletcher, J.S.; Johnson, F.L.; McFarlane, J.C. 1988. Database

assessment of phytotoxicity and data published on terrestrial

vascular plants. Environmental Toxicology & Chemistry. 7:

615–622.

Freeman, D.H.; Angeles, R.M.; Poinescu, I.G. 1976. High

performance liquid chromatography separation of RDX and KMX

explosives on adsorptive polymers. Journal of Chromatography.

118: 157–166.

Gregus, Z.; Klaassen, C.D. 1996. Mechanisms of toxicity. In:

Klaassen, C.D.; Amdur, M.O.; Doull, J.D., eds. Casarett and

Doull’s toxicology: the science of poisons. New York: McGraw-Hill

Publishers: 35–74.

Jenkins, T.F. 1989. Development of an analytical method for

determination of extractable nitroaromatics in soils. University of

Maryland. U.M.I. Dissertation Services. Bell and Howell Co. Ann

Arbor, MI. Ph.D. Dissertation.
Holland, Marjorie M.; Warren, Melvin L.; Stanturf, John A., eds. 2002. Proceedings of a

conference on sustainability of wetlands and water resources: how well can riverine

wetlands continue to support society into the 21st century? Gen. Tech. Rep. SRS–50.

Asheville, NC: U.S. Department of Agriculture, Forest Service, Southern Research Station.

191 p.

Abstract: The conference focused on recent work in freshwater wetlands [both natural and

constructed] with a view toward understanding wetland processes in a watershed context.

Since humans have played important roles in watershed dynamics for years, attention was

given to the human dimensions of wetland and watershed uses. Contributed sessions were

organized on: biogeochemical cycling in wetlands; human health issues related to water;

wetland restoration and reforestation; the role of wetlands in agricultural systems; wetlands

and USA environmental law; chemical ecology and natural products from wetlands; water and

wetlands in science education; and regional water strategies. The lead paper in the

proceedings was prepared by conference plenary speaker Dr. Sandra Postel, director of the

Global Water Policy Project in Amherst, MA. Examples and experiences from eight countries

were shared during the conference, providing valuable global perspectives.

Keywords: Environmental law, fresh water, restoration, sustainable systems, water

resources, watersheds, wetlands.
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