V O L U M E 56
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Advisory Board Martin Alexander Cornell University
Eugene J. Kamprath North Carolina State Univers...
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V O L U M E 56
$
Advisory Board Martin Alexander Cornell University
Eugene J. Kamprath North Carolina State University
Kenneth J. Frey
Larry P. Wilding
Iowa State University
Texas A&M University
Prepared in cooperation with the
American Society of Agronomy Monographs Committee P. S. Baenziger J. Bartels J. N. Bigham L. P. Bush
M . A. Tabatabai, Chairman R. N. Carrow W. T. Frankenberger, Jr. D. M. Kral S. E. Lingle
G. A . Peterson D. E. Rolston D. E. Stott J. W. Stucki
D V A N C E S I N
onomy V O L U M5 6E Edited by
Donald L. Sparks Department of Plant and Soil Sciences University of Delaware Newark, Delaware
W
ACADEMIC PRESS San Diego New York Boston London Sydney Tokyo Toronto
This book is printed on acid-free paper.@ Copyright 0 1996 by ACADEMIC PRESS, INC. All Rights Reserved. No part of this publication may be reproduced or transmitted in any form or by any means, electronic or mechanical, including photocopy, recording, or any information storage and retrieval system, without permission in writing from the publisher.
Academic Press, Inc. A Division of Harcourt Brace & Company 525 R Street, Suite 1900, San Diego, California 92101-4495 United Kingdom Edition published by Academic Press Limited 24-28 Oval Road, London N W I 7DX
International Standard Serial Number: 0065-2 I 13 International Standard Book Number: 0-12-000756-8 PRINTED IN THE UNITED STATES OF AMERICA 96 97 9 8 9 9 00 01 BB 9 8 7 6 5
4
3 2
1
Contents ..............................................
vii
PREFACE ...................................................
ix
CONrRIBUrOKS
SOILHEALTH AND SUSTAINABILITY J . W. Doran. M . Sarrantonio. and M . A. Liebig I . Overview . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . II . Soil - A Vital. Living. and Finite Kesource ............................ I11. Early Proponents of Soil Health Concepts . . . . Iv. Soil Health and Human Health . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . V. Agriculturc and Soil Health . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . VI. Assessment of Soil Quality and Health ................. VII. Soil Assessment - Need for Producer/Scientist Interaction .............. ................ VIII . Summary and Conclusions References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
1
3 11 14
20 28 39 44 45
PHYTOREMEDIATION OF SOILSCONTAMKNATED WITH ORGANIC POLLUTANTS Scott D . Cunningham. Todd A. Anderson. A . Paul Schwab. and F. C. Hsu 1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ...... ................... I1 . “Phytoremediation” . ...... ................... I11. Xenobiotics in Soil .. N . Plants as Kemediation Structure for Organics .... . . . . . . . . . . . . . . . . . . . . . .... .... ...... V. Phytoreniediation ex Plmta VI. Modeling Phytoremediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ........................... VII. Practical Considerations
VIII . Current Phytoremediation Research and Development . Lx. Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . ....
56 61 67
71 82 91
92 99 107 107
BIOLOGICALCONTROL OF WEEDSWITH PLANT PATHOGENS AND MICROBIAL PESTICIDES David 0. TeBeest I. I1 . I11.
Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Strategies for the Control o f Weeds with Plant Pathogens . . . . . . . . . . . . . . . Biological Control of Weeds with Plant Pathogens . . . . . . . . . . . . . . . . . . . . .
V
115 116
117
vi
CONTENTS
IV Biological Control of Weeds by Microbial Management of Seed Banks . . . . V. Synergisms That May Affect the Effectiveness of Microbial Agents . . . . . . . VI . The Environmental Impact of Microbial Herbicides . . . . . . . . . . . . . . . . . . .
VII . Summary ........................................................ References .......................................................
125 125 129 131 132
ORGANIC AMENDMENTS AND PHOSPHORUS SORPTION BY SOILS F. Iyamuremye and R . P. Dick
I. I1.
111.
IV. V.
Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Aerobic Soils: Organic Acids and Phosphorus Sorption ................. Aerobic Soils: Plant Residues and Animal Manures ..................... Waterlogged Soils . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Research Needs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References .......................................................
139 146
156 167 176 178
ADVANCES m DROUGHT TOLERANCE IN PLANTS John S . Boyer
..................... I . Introduction ........................... I1. Water Use Efficiency ................................ ............................... 111. .......... N . Water Deficits and Reproduction ....
v.
VI. Conclusions ........................
References .........................
...................
................... ..............
187 188 196 204 207 210 212
THE AFLATOXINPROBLEM WITH CORNGRAIN Neil W. Widstrom .... ................................ I . Introduction . . . . . tion of Aflaroxins as Contaminants of Corn .... I1. 111. Conditions Impacting Asperg and Aflatoxin Accumulation ............................... owth and Ear Development . . . . . . . N . Managing Conditions during V. Handling the Grain Crop a t Harvest ................................. VI . Storage and Utilization of the Final Product .......................... .................... VII . Long-Range Solutions . . . . . . . . . . . . . . . . VIII . Conclusions., . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References ..... .................................
220 220
I N I ~ E X.....................................................
281
.
226 236 247 249 256 260 261
Contributors Numl)crs in p~rcnthcscsindicatc thc pages on which the authors’ contriburlons t q i n
TODD A. ANDERSON (59, Pesticide Toxicology Laboratory, Iowa State University, Ames, Iowa 5001 1 JOHN S. B O E R (187), College ofAyicuiture and Marine Studies, University of Delaware, Lemes, Delnwai-e 19958 SCOTT D. CUNNINGHAM (5 5), Du Pont Environmental Biotechnology, Glasgow Site, Newark, Delaware 19714 R. P. DICK (1 39), Department of Crop and Soil Science, Oregon State University, Cornallis, Oregon 9 7331 J. W. DORAN (I), Soil and W4ter Conservation Research Unit, United States Depnr-t.ment of Agriculture, Agrirultumi Research Senice, University of Nebraska, Lincoln, Nebraska 68583 F. C . HSU (SS), Du Pont Envir-onnrental Biotechnology, Glasgow Site, Newark, Delaware 19714 F. IYAMUREMYE (1 39), Department of Crop and Soil Science, Oregon State University, Conwllis, Oregon 97331 M. A. LIEBIG (l), Depnrtment of Agronomy, University of Nebraska, Lincoln, Nebraska 68583 M. SARRANTONIO (l), Rodale Institute Research Center, Kutztown, Pennsylvania 19530 A. PAUL SCHWAB (55), Department of Agronomy, Krrnsas Stnte University, Mnnbattan, KNnsas 66506 DAVID 0. TEBEEST (1 15), Department of Pinnt Pathology, Unive~+y of Arkansas, Fayetteville, Arkansas 72701 NEIL W. WIDSTROM (2 19), United States Deparhnent of Agriculture, Agriadturai Research Senice, Georgia Coastal Plain Experiment Station, Tqton, Georgia 3 1793
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Preface Volume 56 contains six cutting-edge reviews on topics that should be of broad interest to crop and soil scientists and, indeed, to professionals in many other fields. The first chapter is a comprehensive review on soil health and sustainability. Subjects that are covered include: an historical perspective on soil health, soil health and its relationship to human health and agricultural sustainability, ways to assess soil quality and health, and integration of soil health concepts into farm management. The second chapter is a state-of-the-art treatise on the use of plants to remediate soils contaminated with organic chemicals. Topics are presented on concepts of phytoremediation, plants as remediation structures for organic pollutants, effects of plant-associated microflora on phytoremediation, and overall advances in phytoremediation research. The third chapter discusses innovative aspects of biological control of weeds with plant pathogens and microbial pesticides. Discussions are included on techniques and strategies, synergisms that can affect biological weed control effectiveness, and environmental impacts of nonchemical approaches. The effect of plant residues, animal manures, and organic acids on the phosphorus chemistry of aerobic and anaerobic soils is fully discussed in in the fourth chapter. Chapter five is a review on advances in drought tolerance in plants. Physiological and molecular biological aspects of water use efficiency are provided as well as the current status of research on drought and desiccation tolerance and water deficits and reproduction. Chapter six deals with the aflatoxin problem in corn grain. Background on the topic, ways to identify aflatoxins, conditions affecting their accumulation, and management regimes and long-term solutions are included. I appreciate the excellent contributions of the authors.
DONALD L. SPARKS
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SOILHEALTH AND SUSTAINABILITY J. W. Doran,' M. Sarrantonio,z and M. A. Liebig3 'SoiVWater Conservation Research Unit, United States Department of Agrkdture, Agricultural Research Service, University of Nebraska, Lincoln. Nebraska 68583 ZRodale Institute Research Center, Kutztown, Pennsylvania 19530 3Deparunent of Agronomy, University of Nebraska, Lincoln, Nebraska 68583
I. Overview 11. Soil - A Vital, Living, and Finite Resource A. Global Function and Sustainability B. Defining Soil Quality and Soil Health 111. Early Proponents of Soil Health Concepts A. Early Scholars and Philosophers B. 19th and 20th Century Scientists and Practitioners W. Soil Health and Human Health A. Direct and Indirect Effects B. Linkages between Soil, Food Quality, and Health V. Agriculture and Soil Health A. Perceptions of Soil B. Regenerative Agriculture C. Natural Resource Accounting VI. Assessment of Soil Quality and Health A. Use of Indicators B. Quantitative Assessments C. Value of Qualitative/Descriptive Assessments VII. Soil Assessment - Need for Producer/Scientist Interaction A. A Shifting Agricultural Research Paradigm B. Integration of Soil Health Concepts into Farm Management C. Technology Transfer VIII. Summary and Conclusions References
I Aduunra in Apnwii.y. Volume Y6 Copyright B 1996 by Academic Press,Inc. MI rights of reproduction in any farm reserved.
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J. W. DORAN ET AL.
I. OVERVIEW Increasing human populations, decreasing resources, social instability, and environmental degradation pose serious threats to the natural processes that sustain the global ecosphere and life on earth (Pearce and Warford, 1993). Agriculture, and society in general, is challenged to develop strategies for sustainability that conserve nonrenewable natural resources such as soil, enhance use of renewable resources, and are aligned with the natural processes that sustain life on earth. The challenge ahead in sustaining life on planet earth will require new vision, holistic approaches for ecosystem management, and a renewed partnership between science and society. We must muster our cultural resources and “put science to work” for both humanity and the natural ecosystems of which it is part and on which it depends. Recznt acceleration of technological growth in industrial and postindustrial societies poses a risk to the health of global ecosystems which are characteristically slow to change (Costanza et a!. , 1992). With the advent of agriculture some 10,000 years ago, the earth’s landscape has been dramatically transformed to yield an abundance of food and fiber to meet the needs of an ever-increasing human population, which increased 600 fold during this time and has twice doubled in size during the past 150 years. For the first 9900 years, agriculture functioned almost entirely on the internal resources available to it from the sun, air, rainfall, plants, animals, soil, and humans and depended on natural processes and ecological associations for its productivity (Rodale, 1995). But about 100 years ago, agriculture began to move beyond its internal resources to production systems based on external inputs such as fertilizer, pesticides, and fossil fuels which had been produced by green plants many millennia in the past. In America, and elsewhere, the achievements of modern science-based agriculture can hardly be overstated. Producers have readily adopted a succession of mechanical, biological, and chemical innovations that have transformed agriculture into a powerful industrial machine that produces abundant food (Northwest Area Foundation, 1994). However, the heavy dependence of modem agriculture on nonrenewable fossil fuels for synthesis of fertilizers and pesticides, and energy needs for cultivation, harvest, intensive animal production, and grain processing, raise questions about the long-term sustainability of agriculture. Also, the full cost of chemical- and energy-intensive agriculture on degradation of natural resources and the quality of air and water environments is rarely debited against gains in productivity (Tangley, 1986). The problems of sustainability which we currently face are considered by some to result from an abandonment of ecological principles to produce human food and the acceptance of a cultural premise that places humankind as the ruler of the world, and therefore not subject to the laws of nature (Quinn, 1993). We often suffer from the delusion that we as
SOIL HEALTH AND SUSTAINABILITY
3
humans can control nature when, in reality, the only thing we can control and manage is ourselves (Cline and Ruark, 1995). The authors of this chapter present the thesis that “soil” is a dynamic, living resource whose condition is vital both to the production of food and fiber and to global balance and ecosystem function, or in essence, to the sustainability of life on earth. The quality and health of soils determine agricultural sustainability (Acton and Gregorich, 1995), environmental quality (Pierzynski et a l ., 1994), and, as a consequence of both, plant, animal, and human health as well (Haberern, 1992). In its broadest sense, soil health can be defined as the ability of soil to perform or function according to its potential, and changes over time due to human use and management or to unusual natural events (Mausbach and Tugel, 1995). In this sense, soil health is enhanced by management and land-use decisions that weigh the multiple functions of soil and is impaired by decisions which focus only on single functions, such as crop productivity. In this chapter we present past and present philosophies of soil health, approaches to assessing the quality and health of soils, and the value of soil health to strategies for sustainable management of our natural resources. Most examples for discussion come from arable agriculture because this is the specialization area with which the authors are most familiar. However, the principles involved apply to forested lands, rangelands, and other terrestrial ecosystems which in some cases may be as or more important to certain aspects of global ecosystem function. The senior author expresses sincere appreciation to coauthors of this chapter for their enthusiastic and valuable contributions and accepts responsibility for any errors in judgment or fact which the chapter may contain.
11. SOIL-A VITAL, LIVING, AND FINITE RESOURCE
A. GLOBAL FUNCTIONAND SUSTAINABILITY We enter the 21st century with greater awareness of our technological capability to influence the global environment and of the impending challenge for sustaining life on earth (Postel, 1994; Gore, 1993). Global climate change, depletion of the protective ozone layer, serious declines in species biodiversity, and degradation and loss of productive agricultural land are among the most pressing concerns associated with our technological search for a higher standard of living for ever-growing human populations. Increasing worldwide concern for sustainable global development and preservation of our soil resources is reflected by numerous recent international conferences such as the United Nations Conference on Environment & Development (UNCED) in Rio de Janeiro, Brazil, in 1992; the Soil Resilience and Sustainable Land Use Symposium in Budapest,
4
J. W. DOKAN ET AL.
Hungary, in 1992; the Sustainable Land Management Conference in Lethbridge, Canada, in 1993; and the International Congress of Soil Science in Acapulco, Mexico, in 1994. Central to discussions at these conferences were the threats to sustainability posed by soil and environmental degradation associated with increasing intensity of land use and the search among increasing populations of the world for a higher standard of living. The sustainability of the energy- and chemically intensive industrial agricultural model, which has enabled a two- to threefold growth in agricultural output of many countries since World War 11, is increasingly questioned by ecologists, earth scientists, and clergy (Jackson and Piper, 1989; Sagan, 1992; Bhagat, 1990). Interest in evaluating the quality and health of our soil resources has been stimulated by increasing awareness that soil is a critically important component of the earth’s biosphere, functioning not only in the production of food and fiber but also in the maintenance of local, regional, and global environmental quality (Glanz, 1995). The thin layer of soil covering the surface of the earth represents the difference between survival and extinction for most land-based life. Like water, soil is a vital natural resource essential to civilization but, unlike water, soil is nonrenewable on a human time scale (Jenny, 1984, 1980). Modem conservationists are quick to point out that “mismanagement and neglect can ruin the fragile resource and become a threat to human survival” (La1 and Pierce, 1991). This is a conclusion supported by archeological evidence suggesting that soil degradation was responsible for extinction or collapse of the Harappan civilization in western India, Mesopotamia in Asia Minor, and the Mayan culture in Central America (Olson, 1981). Present-day agriculture evolved as we sought to control nature to meet the food and fiber needs of an increasingly urbanized society. With the development of modern chemistry during and after World War 11, agriculturists often assumed a position of dominance in their struggle against a seemingly hostile natural environment, often failing to recognize the consequences of management approaches upon long-term productivity and environmental quality. Increased monocultural production of cash grain crops, extensive soil cultivation, and greater reliance on chemical fertilizers and pesticides to maintain crop growth have resulted in twoto threefold increases in grain yields and on-farm labor efficiency (Avery, 1995; Brown et al., 1994; Northwest Area Foundation, 1994; Power and Papendick, 1985). However, in some cases, these management practices have also increased soil organic matter loss, soil erosion, and surface and ground water contamination in the U.S.A. and elsewhere (Gliessman, 1984; Hallberg, 1987; Reganold et af.,1987). Motivations for shifting from input-intensive management to reduced external input farming include concern for protecting soil, human, and animal health from the potential hazards of pesticides, concern for protection of the environment and soil resources, and a need to lower production costs (Soule and Piper, 1992; U.S. Dept. of Agriculture, 1980).
SOIL HEALTH AND SUSTAINABILITY
5
Past management of agricultural and other ecosystems to meet the needs of increasing populations has taxed the resiliency of soil and natural processes to maintain global balances of energy and matter. The quality of many soils in North America has declined significantly since grasslands and forests were converted to arable agriculture and cultivation was initiated (Campbell et a l . , 1976). Mechanical cultivation and the production of continuous row crops has resulted in soil loss through erosion, large decreases in soil organic matter content, and a concomitant release of organic carbon as carbon dioxide to the atmosphere (Houghton et al., 1983). As publicized in the national press, recent inventories of the soil’s productive capacity indicate severe degradation on well over 10% of the earth’s arable land within the last decade as a result of soil erosion, atmospheric pollution, cultivation, over-grazing, land clearing, salinization, and desertification (Sanders, 1992; World Resources Institute, 1992). Findings from a project of the United Nations Environment Program on “Global Assessment of Soil Degradation” indicate that almost 40% of agricultural land has been adversely affected by human-induced soil degradation, and that more than 6% is degraded to such a degree that restoration of its original productive capacity is only possible through major capital investments (Oldeman, 1994). The quality of surface and subsurface water has been jeopardized in many parts of the world by intensive land management practices and the consequent imbalance of C, N, and water cycles in soil. At present, agriculture is considered the most widespread contributor to nonpoint source water pollution in the U.S.A. (CAST, 1992b; U .S. Environmental Protection Agency, 1984; National Research Council, 1989). The major water contaminant in North America and Europe is nitrate-N, the principal sources of which are conversion of native to arable land use, animal manures, and fertilizers. Soil management practices such as tillage, cropping patterns, and pesticide and fertilizer use are known to influence water quality. However, these management practices can also influence atmospheric quality through changes in the soil’s capacity to produce or consume important atmospheric gases such as carbon dioxide, nitrous oxide, and methane (CAST, 1992a; Rolston et al., 1993). The present threat of global climate change and ozone depletion, through elevated levels of atmospheric gases and altered hydrological cycles, necessitates a better understanding of the influence of land management on soil processes. Development of sustainable agricultural management systems has been complicated by the need to consider their utility to humans, their efficiency of resource use, and their ability to maintain a balance with the environment that is favorable both to humans and to most other species (Harwood, 1990). We are challenged to develop management systems that balance the needs and priorities for production of food and fiber with those for a safe and clean environment. In the U.S.A., the importance of soil quality in maintaining balance between environmental and production concerns was reflected by a major conclusion of a
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J. W. DORAN ET AL.
recent National Academy of Science report that “Protecting soil quality, like protecting air and water quality, should be a fundamental goal of national environmental policy” (National Research Council, 1993a). A recent call for development of a “soil health index” was stimulated by the perception that human health and welfare are associated with the quality and health of soils (Haberern, 1992). However, defining and assessing soil quality or health is complicated by the fact that soils perform multiple functions in maintaining productivity and environmental well-being. Identifying and integrating the physical, chemical, and biological soil attributes which define soil functions is the challenge (Papendick and Parr, 1992; Rodale Institute, 1991). Forums were held in Washington, DC, in the winter of 1995 to ensure that emphasis on maintaining the quality of our soil resources was included in the 1995 Farm Bill. Many people recognize that maintaining the health and quality of soil should be a major goal of a “sustainable” society. An important question, however, is “what defines a healthy or quality soil and how might soil quality and health be maintained or improved through agricultural and land-use management?”
B. DEFINING SOILQUALITY AND SOILHEALTH 1. Soil -A Complex Living Ecosystem
Soil forms the thin skin of unconsolidated mineral and organic matter on the earth’s surface and functions to maintain the ecosystems on which all life depends. Soil is a dynamic, living, natural body that is vital to the function of terrestrial ecosystems and represents a unique balance between the living and the dead (Fig. 1). The perception that soil is “living,” though disputed by some, results from the observation that the number of living organisms in a teaspoon of fertile soil (10 g) can exceed nine billion, one and one-half times the human population of the earth. Soils form slowly, averaging 100 to 400 years per centimeter of topsoil, through the interaction of climate, topography, living organisms (microorganisms, animals, plants, and humans), and mineral parent material over time; thus the soil resource is essentially nonrenewable in human life spans (Jenny, 1980; Lal, 1994). Soils are composed of different sized inorganic mineral particles (sand, silt, and clay), reactive and stable forms of organic matter; a myriad of living organisms (earthworms, insects, bacteria, fungi, algae, nematodes, earthworms, etc.), water, and gases including O,, CO,, N,, NO,, and CH,. The physical and chemical attributes of soil regulate soil biological activity and interchanges of molecules/ions between the solid, liquid, and gaseous phases which influence nutrient cycling, plant growth, and decomposition of organic materials. The inorganic components of soil play a major role in retaining cations through ion exchange and nonpolar organic compounds and
SOIL HEALTH AND SUSTNNABILITY
7
Figure 1 A healthy soil is full of macro- and microorganisms in proper balance with the physical and chemical condition of soil (Courtesy of American Journal of Alternative Agriculture, Volume 7 , 1992).
anions through sorption reactions. Essential parts of the global C, N , P, and S and water cycles occur in soil and soil organic matter is a major terrestrial pool for C, N, P, and S; the cycling rate and availability of these elements is continually being altered by soil organisms in their constant search for food and energy sources. The sun is the basis for most life on earth and provides radiant energy for heating the biosphere and for the photosynthetic conversion of carbon dioxide (CO,) and water into food sources and oxygen for consumption by animals and other organisms. Most living organisms utilize oxygen to metabolize these food sources, capture their energy, and recycle heat, CO,, and water to the environment to begin this cycle of life again. A simplified version of this “Equation of Life” can be depicted as follows.
J. W. DORAN ET AL.
8
Photosynthesis KO,
(radiant) Energy (heat)
+ 6H,O +
*
(food) C6H1206
+
602
(fuel) Decomposition & Combustion
The amount of CO, in the atmosphere is rather small and represents less than 0.04% of all gases in the atmosphere. If all the combustion and respiration processes on earth were halted the plant life of the earth would consume all available CO, within a year or two (Lehninger, 1973). Thus, there is a fine balance between CO, production and utilization in the biosphere. Decomposition processes in soil play a predominant role in maintaining this balance. These processes are brought about by a complex web of organisms in soil, each playing unique roles in the physical and chemical breakdown of organic plant and animal residues. The physiological diversity of this decomposer community, however, enables continued activity over a wide range of conditions, an essential attribute in a soil environment which is continually changing. Soils breathe and play a major role in transforming sunlight and stored energy and recycling matter through plants and animals. As such, living soils are vital to providing human food and fiber needs and in maintaining the ecosystems on which all life ultimately depends.
2. The Concept of Soil Quality-Soil Function Blum and Santelises (1994) describe a concept of sustainability and soil resilience based on six main soil functions-three ecological functions and three which are linked to human activity. Ecological functions include biomass production (food, fiber, and energy); the soil as a reactor which filters, buffers, and transforms matter to protect the environment, groundwater, and the food chain from pollution; and soil as a biological habitat and genetic reserve for many plants, animals, and organisms which should be protected from extinction. Functions linked to human activity include the soil as a physical medium, serving as a spatial base for technical and industrial structures and socioeconomic activities such as housing, industrial development, transportation systems, recreation, and refuse disposal; soil as a source of raw materials supplying water, clay, sand, gravel, minerals, etc.; and soil as a cultural heritage, forming part of our cultural heritage, and containing palaentological and archaeological treasures important to preserving the history of earth and humankind. Our concepts of soil quality change as we become aware of the many essential functions soil performs in the biosphere, in addition to serving as a medium for plant growth, and as societal priorities change. In the late seventies, Warkentin and Fletcher (1977) discussed the evolution of soil quality concepts in intensive agriculture. The oldest and most frequently used concept was one of “suitability
SOIL HEALTH AND SUSTAINABILITY
9
for chosen uses,” with emphasis on capability to support crop growth or engineering structures. This evolved to a “range of possible uses” concept which is ecologically based and recognizes the importance of soil to biosphere function and its multiple roles in enhancing biological productivity, abating pollution, and even serving to enhance human health and aesthetic and recreational use of landscapes. Another stage in this evolution was development of the “intrinsic value” concept of soil as a unique and irreplaceable resource, of value apart from its importance to crop growth or ecosystem function. As noted by Warkentin (1995), this view of soils is not widely explored by soil scientists but is held in various forms by naturalists and people who see a special relationship with the earth (Leopold, 1949). Historically soil has been used as an ideal waste disposal system, a biological incinerator destroying all the organic wastes deposited on or in it over time. However, in the 1960s and 1970s it became increasingly apparent that soils were receiving wastes of a type and at a rate that overwhelmed their assimilative capacity, threatened soil function, and called for a major responsibility by agriculturists in defining soil quality criteria (Alexander, 197 1). The quality of soil, as opposed to its health, is largely defined by soil function or use and represents a composite of its physical, chemical, and biological properties that: (i) provide a medium for plant growth and biological activity; (ii) regulate and partition water flow and storage in the environment; and (iii) serve as an environmental buffer in the formation and destruction of environmentally hazardous compounds (Larson and Pierce, 199 1, 1994). Soil serves as a medium for plant growth by providing physical support, water, essential nutrients, and oxygen for roots. The suitability of soil for sustaining plant growth and biological activity is a function of physical properties (porosity, water holding capacity, structure, and tilth) and chemical properties (nutrient supplying ability, pH, salt content, etc.), many of which are a function of soil organic matter content. Soil plays a key role in completing the cycling of major elements required by biological systems ( C , N , P, S , etc.), decomposing organic wastes, and detoxifying certain hazardous compounds. The key role played by soils in recycling organic materials into carbon dioxide and water and degrading synthetic compounds foreign to the soil is brought about by microbial decomposition and chemical reactions. The ability of a soil to store and transmit water is a major factor regulating water availability to plants and transport of environmental pollutants to surface and ground water. Much like air or water, the quality of soil has a profound influence on the health and productivity of any given biome and the environments and ecosystems related to it. However, unlike air or water for which we have quality standards, the definition and quantification of soil quality is complicated by the fact that it is not directly ingested or respired by humans and animals as are air and water. Soil quality is often thought of as an abstract characteristic of soils which cannot be defined because it depends on external factors such as land use and soil manage-
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ment practices, ecosystem and environmental interactions, socioeconomic and political priorities, and so on. Historically, perceptions of what constitutes a “good” soil vary depending on individual priorities for intended soil and land use. However, to manage and maintain our soils in an acceptable state for future generations, soil quality must be defined, and the definition must be broad enough to encompass the many functions of soil. These considerations led to the following definition: Soil quality is the capacity of soil to function, within ecosystem and land-use boundaries, to sustain biological productivity, maintain environmental quality, and promote plant, animal, and human health (after Doran and Parkin, 1994).
3. Defining Soil Health The terms soil quality and soil health are often used interchangeably in the scientific literature and popular press with scientists, in general, prefemng “soil quality” and producers preferring “soil health” (Harris and Bezdicek, 1994). Some prefer the term soil health because it portrays soil as a living, dynamic organism that functions holistically rather than as an inanimate mixture of sand, silt, and clay. Others prefer the term soil quality and descriptors of its innate quantifiable physical, chemical, and biological characteristics. Much discussion at a recent soil health conference in the midwest U.S.A. centered on the importance of defining soil health (Soil Health: The Basis of Current and Future Production, Decatur, IL, December 7, 1994). In those discussions it was observed that efforts to define the concept of soil health have produced a polarization of attitudes concerning the term. On the one hand are those, typically speaking from outside agriculture, who view maintenance of soil health as an absolute moral imperative-critical to our very survival as a species. On the other hand is the attitude, perhaps ironically expressed most adamantly by academics, that the term is a misnomer-a viewpoint seated, in part, in fear that the concept requires value judgments which go beyond scientific or technical fact. The producers, and therefore society’s management of the soil, are caught in the middle of these opposing views and the communication failures that result. “Health” is defined as “the condition of an organism or one of its parts in which it performs its vital functions normally or properly” (Webster’s Third New International Dictionary, 1986). The word is derived from the Old English word haelrh, which was itself derived from the concept of “whole” from hal-whole, healthy-more at whole. Dr. David White, a natural resource economist and speaker at the aforementioned soil health conference, proposed that any definition of soil health should: (i) reflect the soil as a living system; (ii) address all essential functions of soil in the landscape; (iii) compare the condition of a given soil against its own unique potential within climatic, landscape, and vegetation patterns; and (iv) somehow enable meaningful assessment of trends. It is interest-
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ing to note that with some modification, the definition of soil quality presented earlier could serve as a definition of soil health. With consideration of the aforementioned factors, soil health can be defined as: the continued capacity of soil to function as a vital living system, within ecosystem and land-use boundaries, to sustain biological productivity, maintain the quality of air and water environments, and promote plant, animal, and human health. The challenge we face, however, is in quantitatively defining the state of soil health and its assessment using measurable properties or parameters. Unlike human health, the magnitude of critical indicators of soil health ranges considerably over dimensions of time and space. For the remainder of this chapter the terms soil quality and soil health will be used synonymously. However, the term soil health is preferred in that it more clearly portrays the idea of soil as a living dynamic organism that functions in a holistic way depending on its condition or state rather than as an inanimate object whose value depends on its innate characteristics and intended use.
111. EARLY PROPONENTS OF SOIL HEALTH CONCEPTS
A. EARLY SCHOLARS AND PHILOSOPHERS Concepts related to soil health have been articulated since ancient times. Roman philosophers were especially aware of the importance of soil to agricultural prosperity, and reflected this awareness in their treatises on farm management. Cato, Varro, Virgil, and Columella stressed the value of soil and promoted agricultural practices that maintained its fertility. Having to work within boundaries of natural fertility, they keenly recognized that many soil attributes were a function of landscape position and parent material, and accordingly recommended cropping practices that would maximize agricultural efficiency. They also offered qualitative criteria for evaluating soil health, with indicators similar to many being used today (Garlynd et a / ., 1994). Though the reasoning used by the philosophers was simple, the principles of farm management espoused in their treatises offer many lessons to current agriculturists: lessons of patience and thoroughness required of an agricultural paradigm based on natural fertility (Harrison, 1913, p. 2). Inherently, fertile soil was held in high regard among the philosophers. When outlining criteria for choosing a farmstead, Cato considered fertile soil to be a primary component: “Take care that you choose a good climate, not subject to destructive storms, and a soil that is naturally strong’’ (after Harrison, 1913, p. 2 I). Varro took this notion further by considering the quality of a farm’s soil to be the deciding factor that determined its worth: ” . . . it is to the nature of the
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soil that we generally allude when we speak of a farm as good or bad’ (after Storr-Best, 1912, p. 28). Maintaining a fertile soil, then, was of paramount importance to the philosophers. Practices suggested to maintain soil fertility included the use of rotations that incorporated green-manuring or legume crops, application of livestock manure to soil, and fallowing. The Georgics of Virgil, translated by Lewis (1940), outlined numerous methods for maintaining soil fertility. Regarding crop rotation and fallowing, Virgil wrote: “So too are the fields rested by a rotation of crops, and unploughed land in the meanwhile promises to repay you” (Book 1, I. 8283). On using livestock manure, he noted: “Whatever plantations you’re setting down on your land, spread rich dung and be careful to cover with plenty of earth” (Book 11, 1. 346-347). Sensitivity to soil characteristics was evident in the cropping practices advocated by the philosophers. Cropping to the character of the land was the rule, not the exception. This belief was expressed by Varro when he wrote: . . . the same soil is not equally suited for all kinds of produce . . . for it is better to plant crops that do not need much nutriment on thinner soil” (after Storr-Best, 1912, p. 28, 63). Cropping to specific soils was suggested by both Cat0 and Varro. Cato, in De Agriculturu, wrote: “Where the soil is rich and fertile, without shade, there the corn-land ought to be. Where the land lies low, plant rape, millet, and panic grass” (after Harrison, 1913, p. 42). Using senses of sight, taste, touch, and smell, the philosophers set down qualitative guidelines for evaluating soil and its suitability to promote growth of particular crops. Soil color was used often in their treatises as an indicator of productivity, with black soils considered the most productive and suitable for corn production. Saline or acid soils were identified by a simple taste test recommended by Virgil: “The taste of fresh water strained through sour soil will twist awry the taster’s face” (after Lewis, 1940, Book 11, 1. 246-247). The soil’s physical condition was considered an important component for successful crop production. In his classification of farmland, Varro found crumbling soils of medium texture to be ideal for farming: . . . the kind of land which will repay cultivation . , . easily crumbles when dug, and neither resembles ashes in texture, nor is very heavy” (after Storr-Best, 1912, p. 36). Similarly, Columella classified “rich and mellow” soils best for crops and pasture (after Simonson, 1968). Pliny used his sense of smell to test soil. He considered the musty odor of freshly plowed soil to be the most telling assessment of a soil’s quality: “It is the odor which the earth, when turned up, ought to emit, and when once found, can never deceive any person: and this will be found the best criterion for judging the quality of the soil’’ (after Harrison, 1913, p. 91). Interestingly, this same criterion is currently being considered by the USDA National Soil Tilth Laboratory for use as a potential indicator of soil health (T. Parkin, 1995, personal communication). ”
”
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l ! h H AND
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C E N T U R Y SCIENTISTS AND PRACTITIONERS
The nineteenth century brought widespread concern over a potential food crisis caused by a rapid increase in human population. As the need to increase food production was apparent, chemists sought to understand better relationships between soils and plants. Initial work focused on the concept that plants fed directly on soil humus. This theory, put forth by Wallerius in the middle of the eighteenth century, was developed further during the first half of the nineteenth century by Thaer and von Wullfen (Usher, 1923). They believed organic matter in soils had to be kept at or near original levels to maintain fertility and avoid reductions in crop yield. Humus, therefore, was considered a primary indicator of soil quality. Research by these scientists indicated levels of soil humus to decrease under cultivation. This finding resulted in predictions that, without additions of organic matter, soils in central Europe would quickly be exhausted causing significant declines in crop yield (Usher, 1923). The humus concept, though profoundly important for its time, was considered simplistic and limited in scope because of its theoretical basis in phlogiston chemistry (Krohn and Schafer, 1983). Among its foremost critics was Justus von Liebig. Liebig acknowledged the importance of hunius as a critical component of soil fertility, but claimed that a number of key elements were essential for plant nutrition instead. Relying on methodological advances in organic elementary analysis, Liebig found plant nutrient requirements could be estimated by analyzing the elemental concentrations in plants and soils and striking a balance between the amounts in the soil and those in the growing plant. Liebig’s thesis centered on the concept that maintenance of soil quality for growth of plants required the establishment of natural, unbroken cycles of essential plant nutrients within the soil. These cycles, however, were perceived as nonexistent in agricultural practices of the time. According to Liebig, the nutritionally extractive characteristics of agriculture could only be offset by addition of essential plant nutrients to the soil in the form of artificial fertilizers. By doing this, producers could claim to develop a nonexploitative relation to nature “like a wave motion within a cycle” (Liebig, 1862, after Krohn and Schafer, 1983). This new paradigm of plant nutrition caught on rapidly and by the turn of the twentieth century, agriculture had evolved into a major production industry. Under this method of agriculture, soil had acquired the status of a “nutrient bin” for plant roots (Simonson, 1968). In opposition to this form of agriculture was a group of scientists and farmers of “privilege” who regarded soil as a living resource. Sir Albert Howard, J. I. Rodale, Lady Eve Balfour, and William Albrecht represented a handful of individuals who believed soil vitality (i.e., soil life) to be a fundamental component of successful and socially responsible agriculture. By their standard, soil was a form of biological capital: capital that could
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be used wisely by adoption of agricultural practices that relied on balanced natural fertility, or unwisely through continued use of practices that relied on external inputs of artificial fertility. They accordingly held the view that the health and prosperity of society depended upon the condition of the soil. Agricultural systems that promoted soil vitality were strongly advocated by this group. In their view, soil vitality was achieved by maintaining a balance of growth and decay in the soil. This balance was considered to be absent in conventional agricultural systems as a result of a disproportionate emphasis on production (Howard, 1943). Sustainable agricultural systems were regarded as balanced by relying upon vast natural reserves of decaying material. In terms of agricultural management, this implied replenishing organic and mineral matter in the soil. Application of compost to soil was generally accepted as the primary method to maintain soil organic matter. J. I. Rodale, in Pay Dirt (1945), outlined 36 advantages of using compost, 15 of which were directly related to improving soil health. Rodale strongly believed the value of compost could not be estimated by chemical composition alone. In his view, the greatest value of compost was in its potential to improve the biological and physical condition of the soil. Although emphasized less than organic matter application, addition of mineral constituents to the soil was encouraged. Howard regarded the success of Hunzan agriculture to be partly due to the silt-size glacial material found in the irrigation water (Howard, 1947, p. 177). Albrecht and Rodale both stressed the importance of renewing the soil mineral fraction by suggesting the application of lime, wood ash, and even rocks to soil. Primary to the philosophy of this group was the belief that soil quality impacted plant, animal, and human health. Diet was considered to be the primary determinant of good health, and nutrition for all terrestrial organisms began “from the ground up” (Albrecht, 1975). So strong was this belief that they claimed soil quality to be an important element of public health. Lady Eve Balfour, in The Living Soil (1948), declared issues of soil management and public health to be inseparable. In fact, she proposed that agriculture should be looked upon as one of the health services, if not the primary health service. Attainment of this status, however, depended on the need to clearly identify a relationship between soil quality and public health using rigorous scientific methods; a difficult or impossible task.
IV. SOIL HEALTH AND HUMAN HEALTH For much of modern agricultural history, the value of new farming techniques and products was judged primarily, if not solely, on their ability to increase food
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production. As discussed earlier, warnings of potential environmental damage associated with modem agriculture were largely unheeded until recent decades. The concept that the method of food production can have an additional direct impact on animal and human health has recently developed, but only tentatively in scientific circles. The proposal that any definition of soil quality or soil health needs to incorporate the soil’s effect on human health as a component of equal importance with productivity and environmental impact was perhaps first publicly articulated at the Conference on Assessment and Monitoring of Soil Quality held at the Rodale Institute, Emmaus, Pennsylvania in July, 1991 (Papendick and Parr, 1992; Rodale, 1991). Little headway has been made since then in defining the indicators of soil quality and associated effects on human health.
A. DIRECT AND INDIRECTEFFECTS There are three general avenues through which the soil may interact with and affect the health of higher animals. First, there is the potential for direct poisoning of animals and people from contaminated soils. This is most likely to be highly localized and may be the result of industrial accidents or improper use or disposal of agrochemicals, industrial chemicals, or radioactive waste products. While the seriousness of such toxic encounters with the soil is not to be taken lightly, the likelihood of the general population being exposed to soils so highly contaminated as to seriously affect health is very small, There are numerous well-documented occurrences of pesticide poisoning (Hodges and Scofield, 1983; Culliney et al., 1992), but most acute farm chemical poisonings occur before the chemicals are applied to the soil, generally during mixing, or during the spray process itself when chemicals are air-borne (Soule and Piper, 1992; NCAMP, 1990). Recent dramatic increases in certain fungal diseases, often fatal, seen in patients suffering from immunodeficiency diseases such as AIDS can be traced to soil origins (Sternberg, 1994). Although naturally occurring, and not normally associated with unhealthy soil conditions, it appears that soil disturbances, whether natural, as from earthquakes, or human initiated, create the conditions necessary for the spores to be propelled into the atmosphere in numbers sufficiently high to infect the human population. A second, more widespread degree of interaction between soil health and animal/human health occurs indirectly, through the soil’s influence on the quality of water and air. It is well-recognized that there are serious public health concerns related to contaminated groundwater, streams, and other surface water supplies, occasionally including acute toxicity, but more often associated with development of cancer and other long-term debilitating diseases. Nitrate in drinking water can cause the potentially fatal methemoglobinemia, or blue baby syndrome, but can also have more insidious carcinogenic effects if transformed
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in the body to nitrosamines (Clancy, 1986). Nitrosamines can also form from the reaction of nitrate with atrazine, a herbicide commonly found in wells in corngrowing regions (Culliney et al., 1992). According to a U.S. Geological Survey assessment, nitrate-N levels in groundwater were above the 10 mg liter-’ limit determined to be safe by the Environmental Protection Agency in at least 25% of sampled wells in 87 counties in the United States, mostly in the Midwest (USDA, 1987). The contamination is generally thought to be a result of inefficient nitrogen management associated with crop and livestock production. The USDA estimates that nearly half the counties in the United States have groundwater supplies vulnerable to pesticide and nitrate contamination, potentially affecting 54 million people who rely on these sources for drinking water (National Research Council, 1989). Air quality can be equally devastated by poor agricultural practice. The combination of dry weather and poor soil management that caused the Dust Bowl of the 1930s created “billowing red-brown clouds that eclipsed the sun and obliterated fences and covered houses and choked animals and people” (Hillel, 1991), degrading air quality thousands of miles away in New York. Although airborne soil pollution on the scale of the Dust Bowl is rare, localized dust storms and tillage-induced soil clouds due to poor soil conservation methods continue to impair air quality and affect those with respiratory disorders worldwide. More catastrophic occurrences related to poor soil quality include landslides, floods, and fires due to deforestation. Desertification occurring in Sahelian Africa and other places has placed soil management practices in the path of a daily life and death struggle against starvation. The third avenue of impact of soil on animal and human health is also indirect, and occurs through the quality of food plants grown on the soil. The effect may be due to the presence of antiquality factors, such as toxic metals (lead or cadmium), pesticides and animal diseases, or through decreased or imbalanced content of necessary plant nutritional compounds, such as vitamins, proteins, and minerals. Of the two categories, the presence of antiquality factors is easier to detect and trace to specific soil factors. Fruit and vegetables marketed in the United States are mandated to be routinely screened for the presence of an array of pesticides, but testing is random, and many pesticides are not detectable by commonly used analytical methods (National Research Council, 1989). As of 1984, a National Resource Council study estimated that only 10% of the ingredients in pesticides had been thoroughly assessed for health effects. The most acute danger from pesticide residues in food occurs when they appear in their original form, having been sprayed directly on the produce, rather than being filtered first through the soil medium. Dangers from soil-borne pesticides are far less apparent, as many soil-applied pesticides are at least partially decomposed by soil organisms within a short period of time. Nevertheless, several crops, such as potatoes, are among
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those considered to be in the high-risk category for cancer due to soil-applied herbicides (Clancy, 1986). Additionally, metabolites from partially decomposed soil-applied pesticides may persist in forms not tested for and whose biological effects are unknown. Other pesticides will persist in their original form for many years; some of these, while not particularly toxic at the original levels of application, may become concentrated in the food chain over time if they are stored in fatty tissue. The health hazard from such bioaccumulation of agricultural pesticides remains as yet largely unquantified (Culliney et a / . , 1992). Heavy metal and toxic element contamination, on the other hand, is generally more identifiable as a soil quality problem. Such problems may result from geological factors, such as high natural occurrence of the elements of interest in bedrock, or be related to poor agricultural management (Allaway, 1975). The Occurrence of selenium in soils demonstrates this point well. The U.S. Plant, Soil, and Nutrition Research Lab in lthaca, New York, has carefully mapped selenium concentrations in soils throughout the United States and has found that areas considered to have selenium levels below optimum for plant growth coincided with areas of high rates of lung, breast, rectal, bladder, esophageal and cervical cancer, although no direct causal link has been established. Low selenium in these soils might be considered a human antiquality factor, despite the fact that it is due to the natural geology of the regions in which it is found. Areas where selenium occurs in toxic levels, such as the Kesterson Reservoir in California, on the other hand, can be highly localized and generally associated with improper water and soil management (Reisner, 1987). Grass tetany, a disease of ruminants associated with magnesium deficiency and possible calcium deficiency, may also be due to low natural occurrence of the minerals, but is often associated with over-fertilization with potassium and/or ammonium fertilizers (Wilkinson and Stuedmann, 1979). Free nitrate can occur as an antinutritive factor in food plants. Nitrate ingested in plant tissue can react in the body as it does when dissolved in drinking water, possibly leading to methemoglobinemia or conversion to carcinogenic nitrosamines. High nitrate content, a problem particularly in leafy greens such as lettuce and spinach, has been linked additionally to reduced protein quality and lowered vitamin contents of food crops (Knorr and Vogtman, 1983; Linder, 1985; Leclerc et a / . , 1991). Several studies report that vegetables grown with biological sources of nitrogen showed significantly lower excess nitrate than those grown under chemical fertility regimes (Ahrens el a / ., 1983; Lairon et al., 1984; Vogtman el a/., 1984; Termine at al., 1987), but caution must be exercised when determining whether one source of fertility is superior to another. While conventional systems tend to provide large quantities of N in a highly soluble form, which may lead to excess N uptake, systems which include spring plowdown of high-N green manure crops can lead to a similar situation under certain conditions (Sarrantonio and Scott, 1988; Doran and Smith, 1991; Campbell et
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al., 1994). Biologically derived N sources may lead to reduced nitrate concentrations in plants because of more gradual, microbially mediated release of soluble N, but it is probable that any cropping system that adequately synchronizes N availability with plant demand will less likely lead to excess nitrate in the crop plant. Soil health assessment should include monitoring of nitrate synchrony with crop needs throughout the year. Aside from antinutritional factors, the quality of food is more difficult to attribute definitively to soil health alone. Hornick (1992) addresses the problem by classifying the numerous factors that affect food quality under the following broad categories: crop plant and variety selection, management, postharvest handling and storage, climate, and soil. Crop species differ markedly in their nutritional needs and ability to absorb nutrients from the soil, and there are tremendous varietal differences even within crop species. Lantz et al. (1958) reported that varieties of dried beans (Phaseolus vulgaris) differed by as much 70% in protein content by variety and location. Cserni and Prohaska (1987) found that nitrate in carrots of different varieties grown under identical conditions ranged from 156 to 270 ppm. The architecture and efficiency of the root system of individual plants have much to do with inherent ability to explore the soil volume and take up nutrients. Plant breeding and selection may in fact have a more significant effect on crop nutritional quality than the medium in which the crop is grown (Clancy, 1986). Additionally, irrigation, weed control, crop maturity at harvest, and postharvest handling all significantly affect crop nutritional quality in ways totally unrelated to the health of the soil (Kader, 1987). Climate can obviously affect food quality directly in terms of plant stress, but has numerous interactions with the soil as well. Soils with higher water holding capacity, for instance, have a greater buffering capacity against drought, which may help maintain crop quality through prolonged dry periods.
B. LINKAGES BETWEEN SOIL,FOODQUALITY, AND HEALTH The connection between soil health and food quality is not entirely straightforward. While soil fertility can have a profound effect on both crop quality and quantity, crop plants can grow and yield well in soils supplied with inorganic plant nutrients which have little int,eraction with the soil medium. Proponents of food grown under production systems geared toward improving soil health, such as organic or biodynamic systems, may feel strongly that such food is nutritionally superior, but rigorous scientific evidence to support this belief has been difficult to obtain. In addition to the previously mentioned studies which reported lowered nitrates in organically grown food, other studies have indicated several different desirable food qualities associated with organic production, including increased vitamin contents (Leclerc et a / . , 1991), increased dry matter (DeElI
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and Prange, 1992), superior storage quality (Petterson, 1977; Knorr and Vogtman, 1983), and higher protein quality, as measured by EAA (essential amino acid) indices (Eppendorfer, 1978). A nearly equal number of studies, however, indicate that farming method per se had little effect on food quality (McSheehy, 1977; Nilson, 1979; Hansen, 1981). The USDA report on organic farming (1980) was unable to verify that organically produced food was nutritionally superior to conventionally grown food. Knorr and Vogtman (1983) outline the problems involved in sifting through the volumes of available evidence on either side of the issue. They point out that such studies are not closely linked to soil health indicators, but are often pot experiments which test the effects of chemical fertilizers vs organically derived fertilizers, as opposed to stabilized organic systems. Additional problems include results reported in fresh weight bases, which may underestimate nutritional value of crops with varying water contents, and failure to test for trace minerals and vitamins, which may constitute subtle but nutritionally significant differences in foods. The failure to link food quality to actual soil health conditions, regardless of method of production, will continue to impede an informed discussion on the relationship between soil health and human nutrition. Nutritional studies seeking to relate food quality to soil quality/health are complicated by the fact that human populations rarely eat food produced from a localized source. Experiments to test the effect of food supply grown under different management systems are hampered by the nearly insurmountable logistical difficulties of controlling food intake in test groups long enough to show significant differences in health. Nutritional studies additionally suffer from the inability to make valid comparisons where food intake levels among test subjects are unequal. The relevance of such studies is also subject to doubt given that few individuals are likely to ever follow the prescribed diet of the test subjects in detail. Animal feeding studies offer some opportunities for studying, under controlled conditions, the effect of food grown under varying soil management schemes, but such studies have been scarce. Work by Velimirov et a / . (1992) reports the findings of a rigorous nutritional study on three generations of rats fed on biologically produced compared to conventionally produced food. While they found no differences in the number of offspring between the two groups, there were fewer perinatally dead offspring in the groups fed biologically produced food, and the mothers in that group had significantly higher weight gains during and after lactation. In a study performed in 1926 (McCarrison), pigeons grew at a faster rate on grains grown with organic fertilizers that those grown with chemical fertilizers. The organically grown grain was thought to have higher vitamin A and B contents, but analytical methods at the time could not entirely substantiate the theory. McSheehy (1977) found that of mice fed diets from grains grown by organic, chemical, or mixed (reduced chemical) farming, those on the mixed
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farming food source had the highest weaning rate, but no other parameters differed significantly. Other studies have shown no apparent differences in animal health or reproductivity related to the method of growing the food (Miller and Dema, 1958). In the short term, human and animal health is far more likely to be affected by gross changes in the types of food eaten-away from high fat foods and toward more vitamin rich ones, for instance-than by food with subtly different nutritional contents related to the way it was grown. Avery (1995) contends that the benefits of chemically intensive agriculture in providing low cost, appealing fruits and vegetables to consumers, thereby increasing their consumption and utility in preventing cancer, far outweigh the small risks associated with the use of chemicals. The presence of toxic residues on food, however, including systemic pesticides persisting in the soil, may in fact prove to be a long-term determinant of human health. In light of this, consumers may choose a preference path that is least likely to provide unpleasant health-related surprises in the future.
V. AGRICULTURE AND SOIL HEALTH A. PERCEPTIONS OF SOIL While early civilizations and practitioners thought of the soil as a nurturing entity (Mother Earth), a life-giver if not a deity (Lal, 1994; Soule and Piper, 1992), modern agricultural science often treats the soil as a physical medium for anchoring plant roots, which can then be bathed in nutrient and growth regulator solutions. It has been well proven that crops can be grown under such management systems, just as they can be grown without soil at all, in hydroponically managed greenhouses. The short-comings of such soil management systems, however, which neglect both the replenishment of organic matter and the maintenance of complex biological communities is readily apparent when one reviews the role of these components in natural ecosystems. As discussed earlier, organic matter is critical in many soils to maintenance of good soil structure, which provides optimal drainage, water-holding capacity, and aeration for crop growth. Organic matter also contributes significantly to cation exchange capacity (CEC), which enables the soil to buffer nutrient concentrations in solution. While hydroponics may grow viable crops in artificially controlled aerated nutrient solutions, large-scale agriculture is simply not feasible in soil lacking good structure and nutrient-buffering capacity. Even on sandy soils, which have little structure and are less vulnerable to structural degradation, production systems that rely on inorganic nutrient supplies and neglect soil organic matter encompass inherent
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inefficiencies. Soils which are incapable of storing nutrients require excessive or continuous addition of soluble nutrients for crop growth, to compensate for losses and inefficiencies. Soluble nutrients in excess of plant and microbial needs will pass beyond the reach of plant roots, with potential consequences to the groundwater already discussed, compounded by the loss of valuable and possibly dwindling nutrient resources. Soil organisms must be acknowledged as key architects in nutrient turnover, organic matter transformation, and physical engineering of soil structure (see Fig. 1). The microbial populations of the soil alone encompass an enormous diversity of bacteria, algae, fungi, protozoa, viruses, and actinomycetes. As many as 10,000 different species may be found in a single gram of soil (Torsvik et al., 1990), just a small sample of the nearly two million species of microorganisms thought to exist worldwide, with a range of form and function beyond current capacity for comprehensive study. While the specific functions and interactions of the majority of these organisms are as yet poorly elucidated, their role as functional groups in soil health regeneration and maintenance is becoming increasingly clear (Kennedy and Papendick, 1995). The microbial biomass is largely responsible for mineralization and turnover of organic substrates (Killham, 1994). It includes both primary and secondary decomposers, aerobic, anaerobic, and switch-hitting digestors, highly specialized consumers of gourmet delicacies and feeding trough generalists, finicky occupants of outlandish environmental niches and highly adaptable opportunists, hard-driven frenzied achievers and slow-metabolizing plodders, diners of rich, fatty substrates and those eking out an existence gnawing on tough lignaceous scrap. As a group, the community of microbial populations acts without regard for the future, but instead responds quickly to favorable conditions, reproducing and consuming with wild abandon until substrate limitations cause population declines, victims of their collective gluttony. They are in turn cannibalized by their surviving compatriots. The result is a continuous cyclic ballet of nutrient uptake and release that enables less ephemeral life forms in the soil to be supplied with their nutritional needs in a somewhat regulated way. The role of larger soil organisms in maintenance of soil quality and health has finally begun to receive much deserved attention in soil science circles with the publication of several excellent review articles in recent years (Berry, 1994; Linden et a l . , 1994; Stork and Eggleton, 1992). Soil fauna cover a range of soil functions beyond that of the soil microbial community. Anderson ( 1988) classifies soil invertebrates into three categories, based primarily on size. Microfauna are those less than 100 p m in diameter and include protozoa, nematodes, and rotifers. They are the aquanauts of the soil, existing in water films around soil particles and free water in soil pores. They function as secondary consumers, feeding largely on bacteria and fungi, thereby speeding the turnover of microbial biomass and their associated nutrients. The diversity in nematode function is
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vast, spanning many different trophic levels, and nematode identification has been suggested as an indicator of soil organism diversity and soil quality (Bohlen and Edwards, 1994; Bongers, 1990; Neher et al., 1995; Parmelee and Alston, 1986). The mesofauna, according to Anderson’s classification, are those invertebrates 100-200 p m in diameter and include mites, Collembola (springtails), and the Enchytraeidae, or pot worms, as well as thousands of species of insects and spiders. They tend to be omnivores, dining on microflora and fauna, as well as other mesofauna and decomposing plant residues. In this way they speed organic matter turnover directly, as well as indirectly, by fragmenting residues, thereby increasing the surface area available for colonization by smaller organisms. Enchytraeidae affect soil structure through creation of aggregates resulting from fecal pellets and through burrowing activities. Macrofauna are greater than 200 pm and include ants and termites as well as the box-office stars of the underworld, the earthworms. Ants and termites can have localized profound effects on soil structure, but earthworms are more ubiquitous and have become unwitting symbols of a healthy, living soil. They can contribute in several ways to soil health. Most notably, earthworm burrows can occupy as much as 1% of the soil volume (Kretzchmar, 1982), aiding in infiltration and flowthrough of water (Lee, 1985), as well as providing pathways for root exploration and faunal habitat. Their feeding habits can help homogenize the topsoil and, in the case of surface feeders, incorporate large amounts of surface litter into deeper soil levels. Their digestive process releases nutrients and fragments of plant residues, leaving behind fertile casts and mucus burrow linings (Berry, 1994). The conditions favoring high earthworm populations overlap to a great degree with those considered indicative of a healthy soil-good soil structure, adequate moisture, sources of fresh organic material, and absence of certain pesticides. There are several studies which in fact show them to be in considerably greater abundance in natural ecosystems than in cultivated land (Barnes and Ellis, 1979; Mackay and Kladivko, 1985) and higher under “sustainable” than conventional management. Unfortunately, their use as an indicator of soil health is complicated by the fact that the conditions which cause them to be absent, or in low numbers, do not correlate entirely with other indicators of soil health. For example, many will burrow into deep soil layers during cold or extended dry periods. Although earthworms are usually present in highly productive soils, some highquality soils may be devoid of earthworm activity due to such factors as tillage or environmental restrictions (Linden et al., 1994). It is becoming increasingly obvious what the consequences of soil organic matter loss are, that soil organisms are both the preservers and the destroyers of soil organic matter, and that human intervention has a profound effect in orchestrating their activities. Clearly, a new vision of the fragile soil resource is needed.
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23
The concept of the soil as a living organism, as discussed earlier, is not new (Balfour, 1948). It is complementary to the Gaia Hypothesis articulated by James Lovelock and Lynn Margulis in the 1960s (Lovelock, 1991) which envisions the whole planet as a living creature, continually manipulating and adjusting existing conditions to favor its own survival. The soil-as-organism model is useful for conceptualizing the various functioning systems in the soil as analogous to animal respiratory, digestive, and circulatory systems. Perhaps a slightly more appropriate paradigm is of the soil as a community. The difference in the community model is that it is largely self-contained-the outputs and waste products of one group become the inputs and energy sources of another. There can then be more complementarity of function than can occur at the single organism level. Within stable communities, there is little loss of nutrients from the system and outflows of water and energy are balanced by inflows, mostly from rainfall and solar radiation. In this context the need for complex, diverse and overlapping functional groups in the soil becomes apparent. There is a need for both generalists, which perform the bulk of everyday chores, and the specialists in the community which fill specialized niches. In such a model, diversity itself may serve as an indicator for soil health. A follow-through of the soil-as-community paradigm is the idea of plant nutrition being more efficient if cycled through a complex web of organisms and their natural environment which is governed by rules that ensure the survival of the whole community. This has been called the “feed the soil, not the crop” tenet of sustainable or regenerative agriculture. Such plant nutrition seeks to mimic natural ecosystems and relies on the yearly mineralization of organic materials by soil microorganisms in response to fluctuating food sources, moisture, aeration, and temperature. To function properly, it requires a continuing commitment to adding sufficient and diverse organic residues, and to maintaining crop rotations that maximize the presence of living roots throughout the year and synchronize nutrient availability closely with crop needs. It demands a more complex management than current conventional agriculture, and necessitates a higher degree of planning, but theoretically will lead to more efficient and environmentally benign nutrient use. A perhaps more profound outcome of a soil that functions as a living community would be the degree of resilience and stability that develops over time. The dynamic combination of diverse-function populations, sufficient energy supplies, and tight nutrient cycling would be expected to provide the basis in agricultural soils for the kind of buffering capacity against environmental stresses seen in equilibrium level natural ecosystems (Hillel, 1991; Soule and Piper, 1992). In a system following this model, shortfalls in yearly nutrient inputs could be supplemented by stored nutrients in the organic matter or microbial populations. The effect of disease and insect invasions would be minimized by a diverse group of antagonistic organisms, and by the presence of a limited proportion of suscepti-
24
J. W. DORAN ET AL.
ble plants present at any given time. At a physical level, years of drought stress would be ameliorated by higher water-holding capacity and more favorable conditions for root growth due to high organic matter, just as the impact of floods would be lessened by good infiltration and drainage. Swift (1994) proposed that assessments of production sustainability should be based on two components-nondeclining crop yield trends, and stability of yield from crop cycle to crop cycle. There is evidence that this sort of stability can in fact be achieved in agricultural systems. After a 5-year transition period, a comparison of conventionally grown crops and organically grown crops showed that all systems had equivalent yields averaged over 9 years, but the organic systems had less year-to-year variation (Hanson et a l . , 1990; Peters, 1994). Dormaar et al. (1988) reported improved tolerance to drought stress in degraded soils receiving animal manures relative to soils receiving only commercial fertilizer. In the Western Corn Belt of the U.S.A., Sahs and Lesoing (1985) found that yields of rainfed corn (Zea mays L.) for organic management systems using animal manures and/or crop rotations were higher than those for conventional monoculture management with fertilizers and pesticides, especially during years of high temperatures and water stress.
B. REGENERATIVE AGRICULTURE While terms for an agriculture that seeks to mimic natural ecosystems are abundant, the term regenerative agriculture, coined by Robert Rodale, is perhaps the most descriptive. Regenerative agricultural theory assumes that food production systems have caused some degradation of the natural resource base, and seeks ways to restore or regenerate them toward their original state through making maximum use of the internal resources available on farms (Rodale, 1984, 1995). The tenets of regenerative agriculture have never been explicitly laid out. However, laying aside the social and economic aspects, in terms of production systems alone, they are essentially the same as those associated with sustainable or “biological” farming, namely: Soil organic matter replenishment is the cornerstone to regenerating soil health. Plant residues are left in the field or returned as compost as much as possible. Animal production systems are designed to return manures to the soil, either directly by pasturing, or by more efficient manure handling and spreading systems. The necessary removal of organic material in the form of harvested crops is compensated for by growing green manure crops or by amending with compost, which may actually be composed of community food waste, thus tightening the nutrient loop. Living cover should be maintained throughout all or most of the year. This
SOIL HEALTH AND SUSTAINABILITY
25
provides plant roots which can take up soluble nutrients throughout the year, further tightening nutrient cycles and decreasing loss. Living cover also protects against erosion, provides habitat and substrate for soil organisms, and increases soil organic residue inputs. Although the feasibility of cover crops may be limited in drier climates by the potential for competition for available water with a grain crop, perennial soil cover is still an ideal to use as a guideline. Diversity is critical at every level. Crops may be grown in polycultures, or in alternating strips, or diversity may be achieved at the whole farm scale, with complex rotations occumng in numerous small fields. Rotations are based on progressions of plants with complementary water and nutrient needs, pest susceptibilities, and root system types. This above-ground diversity may be expected to harbor below-ground diversity in the soil microbial, and faunal communities as there is a greater variety in food and nutrient sources available. Farm animals may also contribute to the diversity, fulfilling various niches in nutrient cycling and waste disposal. Inorganic fertilizers and pesticides should be reduced or eliminated. While inorganic fertilizers may provide nutrients in similar or identical forms as mineralized organic sources, they are discouraged because they have no direct long-term soil enhancing properties. Certain plant nutrients need to be provided in inorganic form to restore losses from crop removal; in such cases naturally occurring minerals are preferred because they can be applied in less concentrated slow-release forms and commonly require less nonrenewable energy for production and distribution. Pesticide reduction has a twofold purpose-to protect farm employees from exposure to harmful substances, and to avoid creating imbalances in communities of soil organisms. Tillage should be minimized. Excessive tillage leads to increases in organic matter decomposition due to physical disruption of aggregates, increased aeration and warming. While some form of soil disturbance may be required to control weeds, less disruptive cultivation implements are favored and multiple strategies for dealing with weed pressure are employed. The theory and the practice of regenerative agriculture are rarely, if ever, entirely meshed, but there are some signs that movement toward these ideals does in fact lead to improvements in soil health. A comparison of organically and conventionally managed tomato agroecosystems in California (Drinkwater et a l . , 1995) showed that soils managed organically for at least 4 years had slightly greater percentages of soil organic matter, lower soluble N concentrations, and higher levels of microbial activity and potentially mineralizable N. A similar study in New Zealand on paired biodynamically managed and conventional farms found higher levels of microbial activity, soil organic matter contents, and soil nutrient supplying capabilities on the biodynamic farms (Reganold et al., 1993). When compared to a continuous grain system, an 8-year agroecological
26
J. W. DORAN ET AL.
rotation in Alberta, Canada, showed evidence of increases in total C, N , and P, available N , P and K, CEC, microbial biomass, and microbial respiration (Wani et al., 1994). The legume-based cropping system in the Rodale farming system trial now exhibits higher organic matter content and microbial biomass (Wander et al., 1994), greater water stable aggregates (Friedman, 1993), and reduced nitrate leaching (Harris et al., 1994) as compared to the conventional system, while maintaining equivalent yields. Other authors have reported improvements in soil characteristics following transition to more complex rotations including legumes (Angers and Mehuys, 1988; Doran and Smith, 1991; Doran and Werner, 1990; Kay et al., 1988), from reducing tillage (Doran and Linn, 1994; Karlen et al., 1989; Angers et al., 1992), or adding organic soil amendments (Dormaar et al., 1988).
C. NATURAL RESOURCEACCOUNTING Current agricultural practices provide an abundant and generally safe supply of food and fiber at an inexpensive cost to the consumer. The cost of agricultural products at the market, however, does not reflect the full cost of the agricultural system. Environmental costs relating to deleterious consequences of contemporary agriculture, such as soil erosion, polluted water supplies, and poisoned wildlife, are currently ignored under conventional agricultural accounting methods. Though these negative consequences may be more an attribute of a larger economic model affecting agriculture than of agriculture itself, the environmental costs are nevertheless transferred from farmers to people in other places or future time periods (Domanico et al., 1986). Estimated environmental costs of agricultural production are significant. Annual off-site damage from soil erosion by water in the United States has been estimated at over $7 billion (Pimentel et al., 1995; Ribaudo, 1989). Damage includes costs associated with the loss of water’s value for recreation, decreased water storage capacity, flooding, dredging ports and navigable rivers, and treating water for industrial and household use. Of the total soil erosion caused by water in the United States, as much as 75% has been attributed to agricultural sources (Pimentel et al., 1976). Wind erosion damage is generally considered to be less severe than that by water, but may be substantial in arid regions. Damage by wind erosion to households and businesses in New Mexico, where two-thirds of the land is used for agriculture, has been estimated to range from $260 to $466 million annually (Piper and Huszar, 1989). Contamination of water by agrichemicals may be the most costly environmental consequence of agricultural production. Annual damage by pesticides and fertilizers to water quality is suspected to range in the billions of dollars (Duda, 1985; Nielsen and Lee, 1987). Costs associated with surface and ground water contamination from agrichemi-
SOIL HEALTH AND SUSTAINABILITY
27
cals include remediation and replacement of contaminated water, impairment of human and animal health, and loss of water fauna and flora (CAST, 1992b; National Research Council, 1993b). Compared to off-site environmental damage, the value of changes in soil health is much more difficult to quantify. This is primarily due to advances in agricultural technology that have masked much of the yield-reducing impact of soil degradation (Crosson, 1982, p. 184). Calculation of nutrient replacement costs from erosion, however, shed some light on the economic magnitude of agriculture’s impact on soil health. Using the approach of Willis and Evans (1977), estimated loss of nitrogen, phosphorus, and potassium in soil eroded by water would amount to over $6 billion annually in the United States. [This calculation assumes an average soil nitrogen, phosphorus, and potassium content of 0.15, 0.12, and 2.2921, respectively. Current fertilizer prices were used for NH,NO,, P20,, and K 2 0 . Soil erosion by water estimated to be 6.9 metric tons per hectare per year on 155 million hectares of cropland (Kellogg et af., 1994).] These costs reflect just a portion of the economic burden that must be incurred by farmers and consumers alike. The costs of soil organic matter loss and soil tilth deterioration are also likely significant, but remain undefined (Bauer and Black, 1994). Given contemporary agriculture’s estimated cost to the environment and soil health, economic consideration of natural resources is clearly necessary to achieve agricultural sustainability. This has motivated scientists to call for the application of natural resource accounting methods to agricultural production (Domanico et a / . , 1986; Tangley, 1986). This call has been addressed through efforts by the World Resources Institute (Faeth rt ul., 1991), who have employed natural resource accounting to incorporate factors of soil health, regional environmental impacts, farm profitability, and governmental policy to evaluate agricultural sustainability. The method used by Faeth et ul. ( I99 I ) to quantify changes in soil health relies upon interconnected ideas of sustainability, business income, and natural resource depreciation. Sustainability implies that economic activity should meet current needs without foreclosing future options (WCED, 1987). Business income encompasses this notion of sustainability when defined as “the maximum consumption in a certain period that does not reduce potential consumption in future periods” (Edwards and Bell, 1961, after Faeth, 1993). By this standard, then, agricultural accounting methods can only be accurate if depreciation in natural resource assets (i.e., soil) is subtracted from net revenues along with the more common forms of farming assets, like machinery and buildings. Faeth et a / . (1991) followed this standard by calculating a soil depreciation allowance in evaluating the economic performance of agricultural production systems. By incorporating output from the Erosion-Productivity Impact Calculator (EPIC) model, the allowance estimated future income losses over a 30-year period from
28
J. W. DORAN E T AL.
the impact of production on the soil resource as declines in crop yield (Williams et al., 1989). Inclusion of the soil depreciation allowance in their evaluation of economic performance resulted in a reduced net farm income of $62 per hectare per year for Pennsylvania’s best conventional corn-soybean management. This cost represents a significant loss of wealth in the natural resource base: a loss, represented by degraded soil health, that is currently ignored by conventional agricultural accounting methods.
VI. ASSESSMENT OF SOIL QUALITY AND HEALTH Establishing an ongoing assessment of the condition and health of our soil resources is vital to maintaining the sustainability of agriculture and civilization. As discussed earlier, the failure of several earlier civilizations was sealed by their disregard for the health of finite soil resources. In today’s energy- and technology-intensive world, the need for maintaining the health of our soil resources is imperative to sustaining productivity for increasing populations and in maintaining global function and balance. Assessment of soil quality and health is invaluable in determining the sustainability of land management systems. A framework for evaluation or an index of soil quality and health is needed to identify problem production areas, to make realistic estimates of food production, to monitor changes in sustainability and environmental quality as related to agricultural management, and to assist government agencies in formulating and evaluating sustainable agricultural and other land-use policies (Acton, 1993; Granatstein and Bezdicek, 1992). Effective identification of appropriate indicators for soil health assessment depends on the ability of any approach to consider the multiple components of soil function, in particular, productivity and environmental well-being. Identification of indicators and assessment approaches is further complicated by the multiplicity of physical, chemical, and biological factors which control biogeochemical processes and their variation in intensity over time and space (Larson and Pierce, 1991). Realistic assessment of soil quality and health, however, requires consideration of the multiple functions of soil and their relative importance as dictated by societal and ecological needs. There is a great need both to determine the status of and to enhance our soil resources. Assessment and monitoring of the quality and health of soils must also provide opportunity to evaluate and redesign soil and land management systems for sustainability. Standards of soil quality and health are needed to determine what is sustainable and what is not, and to determine if soil management systems are functioning at acceptable levels of performance. Recently, Doran and Parkin (1994) identified nine research needs critical to assessment and enhancement of soil quality. The two highest priority needs were: (i) Establishment of reference
SOIL HEALTH AND SUSTAINABILITY
29
guidelines and thresholds for indicators of soil quality that enable identification of relationships between measured soil attributes and soil function which permit valid comparisons across variations in climate, soils, landuse, and management systems; and (ii) development of a practical index for on-site assessment of soil quality and health for use by farmers, researchers, extension, and environmental monitors that can also be used by resource managers and policy makers to determine the sustainability of land management practices.
A. USEOF INDICATORS Assessing the health or quality of soil can be likened to a medical examination for humans where certain measurements are taken as basic indicators of system function (Larson and Pierce, 1991). In a medical exam, the physician takes certain key measurements of body system function such as temperature, blood pressure, pulse rate, and perhaps certain blood or urine chemistries. If these basic health indicators are outside the commonly accepted ranges, more specific tests can be conducted to help identify the cause of the problem and find a solution. For example, excessively high blood pressure may indicate a potential for system failure (death) through stroke or cardiac arrest. The problem of high blood pressure may result from the lifestyle of the individual due to improper diet, lack of exercise, or high stress level. To assess a dietary cause for high blood pressure, the physician may request a secondary blood chemistry test for cholesterol, electrolytes, etc. Assessment of stress level as a causative factor for high blood pressure is less straightforward and generally involves implementing some change in lifestyle followed by periodic monitoring of blood pressure to assess the effectiveness of the change. This is a good example of using a basic indicator both to identify a problem and to monitor the effects of management on the health of a system. Applying this human health analogy to soil health is fairly straightforward. Larson and Pierce (1991) proposed that a minimum data set (MDS) of soil parameters be adopted for assessing the health of world soils, and that standardized methodologies and procedures be established to assess changes in the quality of those factors. A set of basic indicators of soil quality and health has not previously been defined, largely due to difficulty in defining soil quality and health, the wide range over which soil indicators vary in magnitude and importance, and disagreement among scientists and soil and land managers over which basic indicators should be measured. Acton and Padbury (1993) defined soil quality attributes as measurable soil properties that influence the capacity of soil to perform crop production or environmental functions. Soil attributes are useful in defining soil quality criteria and serve as indicators of change in quality. Attributes that are most sensitive to management are most desirable as indicators
30
J. W. DORAN ET AL.
and some such as soil depth, soil organic matter, and electrical conductivity are often affected by soil degradation processes (Arshad and Coen, 1992). To be practical for use by practitioners, extension workers, conservationists, scientists, and policy makers over a range of ecological and socioeconomic situations the set of basic soil quality/health indicators should meet the following suitability criteria: 1. Encompass ecosystem processes and relate to process-oriented modeling. 2. Integrate soil physical, chemical, and biological properties and processes. 3. Be accessible to many users and applicable to field conditions. 4. Be sensitive to variations in management and climate. The indicators should be sensitive enough to reflect the influence of management and climate on long-term changes in soil quality but not be so sensitive as to be influenced by short-term weather patterns. 5 . Where possible, be components of existing soil data bases.
The need for basic soil quality and health indicators is reflected in the question commonly posed by practitioners, researchers, and conservationists: “What measurements should I make to evaluate the effects of management on soil function now and in the future?” Too often scientists confine their interests and efforts to the discipline with which they are most familiar. Microbiologists often limit their studies to soil microbial populations, having little or no regard for soil physical or chemical characteristics which define the limits of activity for microorganisms, plants, and other life forms. Our approach in defining soil quality and health indicators must be holistic, not reductionistic. The indicators chosen must also be measurable by as many people as possible, especially managers of the land, and not limited to a seleci cadre of research scientists. These indicators should define the major ecological processes in soil and ensure that measurements made reflect conditions as they actually exist in the field under a given management system. They should relate to major ecosystem functions such as C and N cycling (Visser and Parkinson, 1992) and be driving variables for processoriented models which emulate ecosystem function. Some indicators, such as soil bulk density, must be measured in the field so that laboratory analyses for soil organic matter and nutrient content can be better related to actual field conditions at time of sampling. Starting with the MDS proposed by Larson and Pierce (1991), we have developed a list of basic soil properties (Table I) which meets many of the aforementioned requirements of indicators for screening soil quality and health. Appropriate use of such indicators, however, will depend to a large extent on how well the relevance of these indicators is interpreted with respect to consideration of the ecosystem of which they are part. Thus, interpretation of the relevance of soil biological indicators apart from soil physical and chemical attributes and their ecological relevance is of little value and, with respect to assessment of soil quality or health, can actually be misleading.
31
SOIL HEALTH AND SUSTAINABILITY Table I Proposed Minimum Data Set of Physical, Chemical, and Biological Indicators for Screening the Condition, Quality, and Health of Soil (after Doran and Parkin, 1994, and Larson and Pierce, 1994)
Indicators of soil condition
Texture
Depth of soil, topsoil, and rooting
Infiltration and soil bulk density (SBD)
Water holding capacity (water retention chardc.)
Soil organic matter (OM) (total organic C and N)
PH
Electrical conductivity
Extractable N. P. and K
Microbial biomass C and N
Relationship to soil condition and function (rationale as a priority measurement)
Physical Retention and transport of water and chemicals; Modeling use, soil erosion and variability estimate Estimate of productivity potential and erosion; normalizes landscape and geographic variability Potential for leaching, productivity, and erosivity; SBD needed to adjust analyses to volumetric basis Related to water retention. transport, and erosivity; available H,O. calculate from SBD,texture, and OM
Ecologically relevant valuesiunits (comparisons for evaluation)
% Sand, silt, and clay; less
eroded sites or landscape positions cm or m; noncultivated sites or varying landscape positions
minl2.5 cm of water and g/cm3; row and/or landscape positions 8 (g/cm’), cm of available
H20130cm: precipitation intensity
Chemical Defines soil fertility. stability, and erosion extent; use in process models and for site normalization Defines biological and cheniical activity thresholds; essential to process modeling Detines plant and microbial activity thresholds; presently lacking in most process models Plant availahle nutrients and potential for N loss; productivity and environmental quality indicators
Compared with upper and lower limits for plant and microbial activity dS/m; compared with upper and lower limits for plant and microbial activity
Biological Microbial catalytic potential and repository for C and N; modeling: Early warning of nianag. effect on OM
kg N or C/ha-30 cm; relative to total C & N or CO, produced
kg C or N I ha-30 cm; noncultivated or native control
kglha-30 cm: seasonal sufticiency levels for crop growth
(cmtinues )
J. W. DORAN E T AL.
32
Table I (continued) ~
~
Indicators of soil condition Potentially mineralizable N (anaerobic incubation)
Soil respiration, water content, and temperature
~
~
~
~
~
Relationship to soil condition and function (rationale as a priority measurement)
Ecologically relevant values/units (comparisons for evaluation)
Soil productivity and N supplying potential; process modeling; (surrogate indicator of biomass) Microbial activity measure (in some cases plants); process modeling; estimate of biomass activity
kg N/ha-30 cni/day; relative to total C or total N contents
kg Clhaiday; relative microbial biomass actvity, C loss vs inputs and total C pool
Data presented in a recent Science magazine article describing soil quality and financial performance of biodynamic and conventional farming management systems in New Zealand are useful in illustrating some of the above-mentioned points (Table 11). Our analyses, however, are not intended as criticisms of this published work as the authors should be commended for their vision in choice of physical, chemical, and biological indicators of soil quality. One point of discussion is the importance of expressing the results of soil quality tests on a volumetric rather than a gravimetric basis and in units for which ecological relevance can be readily ascertained. As illustrated in Table 11, the magnitude of differences in soil C , total N, respiration, and mineralizable N between management systems for samples expressed by weight of soil are 8 to 10% greater than where expressed on a volume basis using soil bulk density estimates. In cultivated systems soil bulk density can vary considerably across the soil surface due to mechanical compaction and throughout the growing season due to reconsolidation of soil after tillage. Soil bulk density is also directly proportional to the mass of any soil component for a given depth of soil sampled. Where samples are taken in the field under management conditions of varying soil densities, comparisons made using gravimetric analyses will err by the difference in soil density at time of sampling. The observed differences due to management in the New Zealand study were statistically significant. However, since results were expressed on a gravimetric basis, they may not be valid or ecologically relevant. Where values for soil bulk density at time of sampling are not available, the use of soil indicator ratios, in this case mineralizable N to C, can reduce errors of interpretation associated with use of results expressed on a weight basis. Reganold and Palmer (1995) recommend calculating soil measurements on a volume basis per unit of topsoil or solum depth for most accurate assessment of management effects on soil quality,
33
SOIL H E b T H AND SUSTAINABILITY Table 11 Reported and Ecologically Relevant Mean Values of Aggregated Soil Quality Data for the 0- to 20-cm Layer of 16 Biodynamic and Conventional Farms in New Zealand (after Reganold el al., 1993)
Soil property
Reported units and values 0-5 cni hulk density (Mg n i - 3 ) Topsoil thickness (cm) Carbon i%) Total N img kg 1 ) Mineralizable N (mg kg 1 ) Respiration (PI 0, h~ I g- 1 ) Ratio: mineralizdble N to C (nig g I ) Extractable P (mg kg-I) PH Ecologically relevant units and values 0-20 cm bulk density" (g a n - ' ) Carbon (Mg ha I ) Total N ( kg N ha-') Mineralizdble N (kg N h a - I l 4 d-I) Respiration in lab ikg C ha- Id I ) Ratio: niineralizahle N to C Extractable P (excess) (kg P ha- ) pH units above 6.0 lower limit
Biodynamic farms
I .07 22.8 4.84 4840 140.0 73.7 2.99 45.7 6.10 I .2 116.2 1 1.616 336 2275 2.89 110 (50)
0. I
Conventional farms 1.15
20.6 4.27 4260 105.9 55.4
2.59 66.2 6.29 I.3 111.0 1 1,076
215 I850 2.48 172 (112) 0.3
Ratio bio./conv.
0.93* 1.11* 1.13*
1.14* 1.32* I .33* 1.15* 0.69* 0.97* 0.92 I .05 1.05 I .22 I .23 1.17* 0.63* 0.33
Estimated, since data were given only for 0-5 cni depth
* Values differ significantly ( p < 0.01).
The choice of units for soil quality indicators can also have an important bearing on determining the ecological relevance of measured values. In the New Zealand study, respiration of laboratory incubated soils from biodynamic farms averaged 73.7 pl 0, h-i g-I, significantly greater (33%) than that from conventional farms. One interpretation of these results could be that the soils of the biodynamic farms are healthier since respiration was greater. However, if one assumes that for aerobic respiration a mole of oxygen is consumed for each mole of carbon dioxide produced, and the results are adjusted for soil density and expressed as kilograms C released per hectare per day, a different picture emerges. The quantities of C released in 1 day from both the biodynamic and conventional farms are incredibly high and represent 2.0 and I .7%, respectively, of the total C pools of these surface soils. While the values for soil respiration from disturbed soils incubated in the laboratory only represent a potential for release of readily metabolizable soil C (labile C), the results clearly demonstrate
34
J. W. DOKAN ET AL.
that more may not be better and these high rates of respiration may be ecologically detrimental as they represent potentials for depletion of soil organic C or accelerated enrichment of the atmosphere with carbon dioxide. When expressed in ecologically relevant units, it becomes obvious that the respiration rates observed in this study are of limited use in evaluating the status of soil quality and health between these different farming management systems. Similar observations can be made for mineralizable N and extractable P. Levels of mineralizable N above that needed for crop production for biodynamic farms and extractable P levels above crop needs for conventional farms could represent a lower level of soil quality and health as a result of greater potential for environmental contamination through leaching, runoff, or volatilization losses. This is another example that, with respect to soil quality and health, more is not necessarily better and ecologically relevant units are needed for proper evaluation. Soil pH is another example of a soil quality attribute that must be referenced to a definable standard for upper and lower limits which are defined by the cropping system or biological processes of greatest ecological relevance. The above discussion serves to highlight the difficulty we have in interpreting results of laboratory incubations and the need for in-field measurements of respiration and N cycling. Indicators of soil quality and health are commonly used to make comparative assessments between agricultural management practices to determine their sustainability. However, the utility of comparative assessments of soil quality are limited because they provide little information about the processes creating the measured condition or performance factors associated with respective management systems (Larson and Pierce, 1994). Also, the mere analysis of soils, no matter how comprehensive or sophisticated, does not provide a measure of soil quality or health unless the parameters are calibrated against designated soil functions (Janzen et a l . , 1992).
B. QUANTITATIVE ASSESSMENTS Quantitative assessments of soil quality and health will require consideration of the many functions that soils perform, their variations in time and space, and opportunities for modification or change. Criteria are needed to evaluate the impact of various practices on the quality of air, soil, water, and food resources. Soil quality and health cannot be defined in terms of a single number, such as the 10 mg liter-' N03-N standard applied for drinking water, although such quantitative standards will be valuable to overall assessment. Assessments must consider not only the specific soil functions being evaluated, but also land use and societal requirements. Threshold values for key indicators must be established with the knowledge that these will vary depending upon land use, the specific soil function of greatest concern, and the ecosystem or landscape within which
35
SOIL HEUTH AND SUSTAINABILITY
the assessment is being made. For example, soil organic matter concentration is frequently cited as a major indicator of soil quality. Threshold values established for highly weathered Ultisol soils in the southeastern United States indicate that surface soil organic matter levels of 2% (1.2% organic C) would be very good, while the same value for Mollisols developed under grass in the Great Plains, which commonly have higher organic matter levels, would represent a degraded condition limiting soil productivity (Fig. 2 ) . As pointed out by Janzen et al. ( 1992) the relationship between soil quality indicators and various soil functions does not always comply to a simple relationship increasing linearly with magnitude of the indicator, as is commonly thought. Simply put, bigger is not necessarily better. Soil quality and health assessments will have to be initiated within the context of societal goals for a specific landscape or ecosystem. Examples include establishing goals such as enhancing water quality, soil productivity, biodiversity, or recreational opportunities. When specific goals have been established or are known, then critical soil functions needed to achieve those goals can be agreed upon, and the criteria for assessing progress toward achieving those goals can be set. Periodic assessments of soil quality and health with known indicators, thresholds, and other criteria for evaluation will then make it possible to assess soil quality and health quantitatively. To accomplish such goals, several approaches for assessing soil quality have been proposed (Acton and Padbury, 1993; Doran and Parkin, 1994; Karlen ct al., 1994; Larson and Pierce, 1994). A common attribute among all these approaches
-.g
8000-
3 V
6000-
8
! b
20001 4000-
'D
b I-
R2 -4.41 0
,
0 0
1
2
3
, , , 4 5 6 Soil organic C (%)
, 7
,
,
0
9
figure 2 Relationship between organic C concentration in the surface 0- 15 cm of soil and soil productivity as determincd by total dry matter yield at dryland site in Alberta, Canada, in 1991 (after Janzcn Pt a / ., 1992; with permission).
36
J. W. DORAN E T AL.
is that soil quality is assessed with respect to specific soil functions. Larson and Pierce (1 994) proposed a dynamic assessment approach in which the dynamics, or change in soil quality, of a management system is used as a measure of its sustainability. They proposed use of a minimum data set of temporally variable soil properties to monitor changes in soil quality over time. They also proposed use of pedotransfer functions (Bouma, 1989) to estimate soil attributes which are too costly to measure and to interrelate soil characteristics in evaluation of soil quality. Simple computer models are used to describe how changes in soil quality indicators impact important functions of soil, such as productivity. An important part of this approach is the use of statistical quality control procedures to assess the performance of a given management system rather than its evaluation by comparison to other systems. This dynamic approach for assessing soil quality permits identification of critical parameters and facilitates corrective actions for sustainable management. Karlen and Stott (1994) presented a framework for evaluating site-specific changes in soil quality. In this approach they define a high quality soil as one that: (i) accommodates water entry, (ii) retains and supplies water to plants, (iii) resists degradation, and (iv) supports plant growth. They described a procedure by which soil quality indicators which quantify these functions are identified, assigned a priority or weight which reflects its relative importance, and scored using a systems engineering approach for a particular soil attribute such as resistance to water erosion. Karlen et al. (1994) also demonstrated the utility of this approach in discriminating changes in soil quality between long-term crop residue and tillage management practices. Doran and Parkin (1994) described a performance-based index of soil quality that could be used to provide an evaluation of soil function with regard to the major issues of (i) sustainable production, (ii) environmental quality, and (iii) human and animal health. They proposed a soil quality index consisting of six elements: SQ = f(SQE1, SQE2, SQE3, SQE4, SQE5, SQE6),
where SQEl = food and fiber production, SQE2 = erosivity, SQE3 = groundwater quality, SQE4 = surface water quality, SQE5 = air quality, and SQE6 = food quality. One advantage of this approach is that soil functions can be assessed based on specific performance criteria established for each element, for a given ecosystem. For example, yield goals for crop production (SQEl), limits for erosion losses (SQE2), concentration limits for chemicals leaching from the rooting zone (SQE3), nutrient, chemical, and sediment loading limits to adjacent surface water systems (SQE4), production and uptake rates for gases that contribute to ozone destruction or the greenhouse effect (SQES), and nutritional composition and chemical residue of food (SQE6). This list of elements is restricted to agricultural situations but other elements could be easily added, such as wildlife habitat quality, to expand the applications of this approach.
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37
This approach would result in soil quality indices computed in a manner analogous to the soil tilth index proposed by Singh et al. (1990). Weighting factors are assigned to each soil quality element, with relative weights of each coefficient being determined by geographical considerations, societal concerns, and economic constraints. For example, in a given region, food production may be the primary concern, and elements such as air quality may be of secondary importance. If such were the case, SQEl would be weighted more heavily than SQE5. Thus this framework has an inherent flexibility in that the precise functional relationship for a given region, or a given field, is determined by the intended use of that area or site, as dictated by geographical and climatic constraints as well as socioeconomic concerns. Assessment of soil quality and health is not limited to areas used for crop production. Forests and forest soils are important to the global C balance as related to C sequestration and atmospheric levels of carbon dioxide. Soil organic matter and soil porosity, as estimated from soil bulk density, have recently been proposed among international groups as major soil quality indicators in forest soils (Richard Cline; personal communication, June 13, 1995). Criteria for evaluating rangeland health have recently been suggested in a National Research Council (1994) report which describes new methods to help classify, inventory, and monitor rangelands. Rangeland health is defined as the degree to which the integrity of the soil and the ecological processes of rangeland ecosystems are sustained. Assessment of rangeland health is based on the evaluation of three criteria: degree of soil stability and watershed function, integrity of nutrient cycles and energy flows, and presence of functioning recovery mechanisms.
C. VALUEOF QUALITATIVE/DESCFUPTWE ASSESSMENTS The concept of soil health is in many ways farmer-generated and rooted in observational field experiences which translate into descriptive properties such as its look, feel, resistance to tillage, and smell. Harris and Bezdicek (1994) conclude that farmer-derived descriptive properties for assessing soil health are valuable for: (i) defining soil qualitylhealth in meaningful terms, (ii) providing a descriptive property of soil quality/health, and (iii) providing a foundation for developing and validating an analytical component of soil health based on quantifiable chemical, physical, and biological properties that can be used as a basis for management and policy decisions. Unfortunately, the potential contributions of indigenous farmer knowledge to management of soil qualitylhealth throughout the world has not been fully utilized (Pawluk et al., 1992). The use of descriptive soil information is not commonly used in scientific literature dealing with characterization of soil quality/health. However, Arshad and Coen (1992) indicate that many soil attributes can be estimated by calibrating qualitative observations against measured values and recommend that qualitative
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(descriptive) information should be an essential part of soil quality monitoring programs. Visual and morphological observations in the field can be used by both producers and scientists to recognize degraded soil quality caused by: (i) loss of organic matter, reduced aggregation, low conductivity, soil crusting and sealing; (ii) water erosion, as indicated by rills, gullies, stones on the surface, exposed roots, uneven topsoil; (iii) wind erosion as indicated by ripple marks, dunes, sand against plant stems, plant damage, dust in air, etc.; (iv) salinization, as indicated by salt crust and salt-tolerant plants; (v) acidification and chemical degradation, as indicated by growth response of acid-tolerant and -intolerant plants and lack of fertilizer response; and (vi) poor drainage and structural deterioration, as indicated by standing water and poor or chlorotic plant stands. Doran et al. (1994a,b) stressed the importance of holistic management approaches which optimize the multiple functions of soil, conserve soil resources, and support strategies for promoting soil quality and health. They proposed use of the basic set of soil quality and health indicators given in Table I to assess soil health in various agricultural management systems. However, while many of these key indicators are extremely useful to specialists (i.e., researchers, consultants, extension staff, and conservationists) many of them are beyond the expertise of the farmer to measure (Hamblin, 1991). In response to this dilemma, Doran (1995) presented strategies for sustainable management which also in-
Table 111 Sustainable Management Strategies for Building Soil Quality and Health and Associated Indicators which Are Assessable by Producers Strategy
Indicators
Conserve soil organic matter (through maintaining balance in C and N cycles where inputs = outputs)
Directionlchange in organic matter levels with time; potential within soil, climate, and cropping patterns; both visual and analytical measures; soil infiltration/water-holding capacity Visual signs (gullies, rills, dust, etc.); surface soil characteristics: depth of topsoil, organic matter content/texture, intiltration rate Crop growth characteristics (yield, N content. color, rooting); soil and water nitrate levels; soil physical condition/compaction; input costs
~~
Minimize soil erosion [through conservation tillage and increased soil cover (residue, cover crops, green fallow, etc.)] Substitution of renewable for nonrenewable resources [through less reliance on synthetic chemicals, conservation tillage, and greater use of natural balance and diversity (crop rotation,legume cover crops, etc.)] Move toward management systems which coexist more with and less dominate natural systems (through optimizing productivity needs with environmental quality)
Crop growth characteristics (yield, N content, color, vigor); soil and water nitrate levels; synchronization of N availability with crop needs during year
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cluded generic indicators of soil quality and health which are measurable by and accessible to producers within the time constraints imposed by their normally hectic and unpredictable management schedules (Table 111).
VII. SOIL ASSESSMENT -NEED FOR PRODUCER/SCIENTIST INTERACTION A. A SHIFTINGAGRICULTURAL RESEARCH PARADIGM Successful integration of soil health concepts into farm management is a monumental task not unlike the soil conservation movement undertaken by Hugh H . Bennett, “father” of the USDA Soil Conservation Service, earlier this century. It will be necessary for public and private agricultural organizations to work together to ensure farmer adoption and legislator approval of management systems that sustain long-term soil productivity. Central to fulfilling this goal is the identification of profitable and environmentally benign management systems that enhance soil quality and health. Understanding how such management systems concurrently achieve these objectives so that they can be easily adopted across different ecoregions is a challenge appropriate for agricultural research. Agricultural research has exclusively addressed problems in agriculture, not the problem of agriculture (Jackson, 1980). This is reflected by a predominant research emphasis on increasing short-term technical and economic efficiency of agricultural production. Though the problem qf agriculture has yet to be addressed, expectations of agricultural research have broadened appreciably in recent years. Expectations now include finding ways to “reduce consumption of non-renewable resources, avoid environmental damage, minimize toxic residues in food, reverse deterioration of rural communities, and, more generally, preserve long-term productive capacity” (Lockeretz and Anderson, 1993, p. 3). These new expectations are primary goals in developing sustainable agriculture (Gardner et a l . , 1995), goals that pose significant challenges to agricultural research. To successfully address these new expectations, agricultural research will likely require integrated, system-level research approaches (Bezdicek and DePhelps, 1994). Unfortunately, the structure of agricultural research makes it poorly suited for this cause (Lockeretz and Anderson, 1993, Chap. 2). Much of agricultural research has followed the more traditional sciences in a disciplineoriented paradigm. This paradigm, developed by Francis Bacon and advanced by Rene Decartes, is based on reductionistic methods that place priority on the parts of things over the whole (Jackson and Piper, 1989). In addition to its obvious inappropriateness for multifaceted research problems, the specialization associ-
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ated with this scientific paradigm has allowed disciplines within agricultural research to become intellectually self-contained. As a result, societal concerns and problems are not always effectively addressed because the “questions and products” of research are determined and reviewed within disciplinary boundaries (Weinberg, 1967). It is not appropriate, however, simply to dismiss scientific advances made by agricultural research following the discipline-oriented paradigm. In fact, information gathered from it is often necessary to solve larger-scale problems. It is appropriate, however, to challenge the validity of this paradigm for all of agricultural research. Agriculture’s impact on the global ecosphere is well established, and therefore, agricultural research possesses significant social purpose. It is in this light of social obligation that many agricultural researchers have recently sought new research strategies to address societal concerns (Bezdicek and DePhelps, 1994; Gardner, 1990; Hendrix, 1987). To meet the new expectations of agricultural research, scientists will likely have to employ alternative research methods such as farmer-participatory research and multidisciplinary cooperation, two research methods generally not utilized by scientists under the current discipline-oriented paradigm. Leading proponents for change in agricultural research have stressed the need for more farmer-participatory research as a means to study innovative management systems, utilize research methods grounded in ecological principles, and increase farmers’ influence over research priorities (Lockeretz and Anderson, 1993, Chaps. 8 and 9). Multidisciplinary approaches have been suggested in farmerparticipatory research, especially in studies evaluating the agronomic or economic performance of whole farms (Bezdicek and DePhelps, 1994). This sentiment has been echoed by farmers critical of the prevalent reductionistic focus in agricultural research (Kirschenmann, 1991 ; Thornley, 1990; Watkins, 1990). They believe that farmer-participatory research would force scientists to view agricultural problems from a farmer’s perspective. Through an appreciation of the interactions and interdependencies within whole farms, they claim scientists would develop a better understanding of the values that motivate farmers’ production decisions and conduct research that more appropriately addressed farmers’ concerns. Should a greater awareness of farmers’ concerns occur in the research community, research questions would likely be directed less toward increasing disciplinary understanding and more toward solving problems. Problem-oriented research, however, would create a dilemma for most agricultural researchers. Unless agricultural researchers could solve farmers’ problems and increase disciplinary understanding, they would run the risk of not faring well professionally as long as peer-reviewed publications were upheld as the standard of achievement (Lockeretz, 1995). In order to survive professionally, problem-oriented researchers would be forced to mold the results of their work to the research
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community, often using the “sterile formalism and jargon of the discipline” (Lockeretz and Anderson, 1993, p. 156). Doing this, however, would almost guarantee that their work would be ignored by the people for whom it was originally intended. This dilemma that leaves researchers unable to wholly address farmers’ problems represents a fundamental flaw in the current agricultural research paradigm. The professional reward system in agricultural research is designed primarily to further discipline understanding, not to solve problems. Tailoring the reward system to the characteristics of alternative research strategies has been suggested as one way to circumvent this flaw (MacRae era/., 1989). Lockeretz and Anderson (1993, Chap. 10) have suggested a more aggressive approach. They believe researchers should think beyond getting the system to accommodate a particular kind of research, and “challenge the very idea of the dominant system as poorly suited to the social purposes of agricultural research.” They propose that the developnient of an appropriate professional reward system would be facilitated by an institutional realignment that divides agricultural departments into farming-related and agricultural science-related research areas. Farming-related research would cover topics closely associated with farms and production systems, while agricultural science-related research would address agriculturally significant processes and organisms abstracted from the context of production. If stated similarly, but by the goals of each area, farming-related research would address farmers’ concerns and agricultural science-related research would answer disciplinary-related questions. A reorganization of this sort would essentially erase current problems in the professional reward system because farmingrelated research would have to use entirely different criteria for evaluation of achievement (Lockeretz, 1995).
B. INTEGRATION OF SOILHEALTH CONCEPTS INTO FARMMANAGEMENT At a time when agriculture must address environmental degradation due to certain yield-promoting practices driven by increasing demands for both greater and better-distributed food supplies, the concept of soil health can be a useful communication device in meeting present and future world needs. Stewardship of the soil resource that enhances soil quality and health while allowing for acceptable long-term production levels is in everyone’s best interest and satisfies what has been called the ‘Ecocentric’ notion of the Common Good (Stauber, 1994). Soil management practices must now be evaluated for their impacts across the temporal scale-short-, middle-, and long-term, as well as across the landscape, to be truly sustainable (Swift et al., 1991). Producers around the globe receive advice, whether provided gratis by govern-
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ment agencies or solicited for a fee from consultants, on recommended production practices. Unfortunately, much of this advice is often aimed at relatively short-term ( 1 or 2 years) economic gains to their operation, rather than on longterm resource conservation (Stauber, 1994). Additionally, advice may be valueladen, or linked to agribusiness sales, such as soil tests performed by private companies which may indicate need for chemical fertilizers and pesticides in excess of what is needed for good crop production (Cramer, 1986; Soule and Piper, 1992). Management recommendations are often developed for regions which may encompass a wide variation in soil type, topography, and resource availability. In such cases, practices which are appropriate for experimental conditions may be inappropriate on a large portion of the individual farms to which they are recommended. To begin the move toward site-specific best management practices, tests for soil quality indicators should be developed as meters for gauging both the short- and long-term effects of various production practices on soil health. Soil quality tests that yield results uncoupled from value judgments will allow both land stewards and researchers to evaluate production practices objectively under a wide range of conditions, to identify those that are truly improving soil health. Clearly, there will likely always be value judgment necessary to reconcile the need for food production with the need to maintain soil in a near-natural state, such as the decision as to whether increasing herbicide use may be an acceptable tradeoff for reducing tillage. Nevertheless, tests which accurately measure the soil quality impacts of various options will help make the consequences of the different options more apparent. If tests are made to be used by producers and other land stewards, production practices will not only be efficiently tailored for individual situations, but researchers will have a manyfold increase in the information available to better understand soil processes. The concept of soil health can be a key tool for educating farmers about some of the less obvious potentials for soil degradation due to poor management. There is some evidence that a concern for soil health may lead land stewards to production practices that indeed improve some soil characteristics. Van Kooten et al. (1990) found in southwestern Saskatchewan that farmer concern for soil quality was in fact correlated to production practices which improved soil physical parameters. The authors found, however, that farmers were less likely to be seriously concerned with soil quality in areas with deep topsoil, which pinpoints the need to emphasize the long-term vision of soil health.
C. TECHNOLOGY TRANSFER Producers and land managers need practical tools which they can use to determine the effectiveness of their management practices on soil health and sustainable production. Traditional research has identified management practices
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that conserve the soil resource, protect air and water quality, or maximize crop yields. However, development of sustainable management strategies that maintain soil quality and health and balance production needs with environmental concerns require new research approaches and on-site evaluation to confirm the specific applications of general strategies across the range of climatic, soil, economic, and social conditions experienced by agriculture. Facilitating producer participation in the research process is essential to development of practical production systems and assessment approaches which address the needs of both producers and society in general. Indicators of soil health and practical assessment tools are essential to forming this necessary partnership between producers and the technical community. National standards of soil quality and health will likely be established within the next decade to provide policy makers and action agencies with a means of monitoring the state of our soil resources. It is imperative that the indicators be useful to producers in some form especially if incentives or regulations based on soil quality or health are enacted. To include producers as active participants in on-site assessment of soil quality/health, tools and methodologies used by researchers must be adapted to be easily accessible to the producers themselves (Sarrantonio et al., 1996). Tests should be simple to perform, require little in the way of expensive equipment, and give rapid results. Additionally, tests should be able to measure soil characteristics that are meaningful to the producers’ understanding of soil and soil processes, and give results that are reliable, accurate within an acceptable range, and interpretable with a minimum amount of training. A soil quality test kit is currently being developed by USDA-ARS to help producers, researchers, conservationists, environmentalists, and consultants assess the health and quality of soil and facilitate technology transfer (Crarner, 1994). The test kit provides on-site capability for assessment of many of the indicators for screening soil quality and health (see Table I ) such as soil pH, electrical conductivity, soil and water nitrate levels, soil density, water infiltration, water-holding capacity, soil water content, water-filled pore space, soil temperature, and soil respiration. The kit provides producers and agricultural specialists with the tools necessary for a cursory assessment of the complex suite of physical, chemical, and biological factors which comprise soil quality/health and facilitates on-site identification of the soil resource condition and its degree of degradation. Currently the cost of the test kit is under $250, yet results obtained with this kit compare well with standard laboratory procedures that are more time consuming and costly (Liebig ct al., 1994). The utility of this test kit is currently being evaluated by conservationists (USDA-NRCS), researchers, extension educators, environmental monitors (EPA-EMAP), and producers at locations in the United States, Australia, Canada, Cuba, Honduras, India, Poland, and Ru . Preliminary results suggest the kit is useful to specialists in fostering appreciation for the complexity of soil, in bridging disciplinary boundaries, and in facilitating assessment of soil quality
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and health. However, the overall procedure for on-site assessment of soil quality and health was found to be too complicated and time consuming for practical use by farmers. One extension educator in Illinois suggested that the “test” kit might best be used by farmers as a ‘tool’ kit from which specific tests can be used as needed to assess soil quality and health. Practical tools for soil quality and health assessment by producers must aid their comprehension of the concept of soil health and be useful to them within the context of their normal work routines (after Nowak in Leopold Letter, 1995). Knowledge of soil for most producers is largely limited to that which they gain through their sensory experiences in working the soil with agricultural implements and watching plant growing conditions during the growing season. Knowledge derived from studying soil test results (mainly organic matter), conservation plans, and information from farm supply dealers are of less importance to farmers in understanding soil. Information from soil conservation offices (USDA-NRCS), taking soil samples, and experience of others are the least relied upon sources of knowledge about soils. Clues farmers most often use to differentiate soils include soil color (largely organic matter), the workability of soil (structure and compaction), wetness or dryness of soil (drainage, storage, and infiltration capacity), and topsoil texture and depth (indicators of soil erosion and production potential). Crop yield and input costs are indicators which producers most often rely upon to assess the short-term sustainability of their management practices. Inclusion of other tools for rapid assessment of efficiency of resource use such as quick tests for soil and water nitrate levels, adequacy of plant growth and N content, and synchronization of soil nitrogen supplies with crop plant needs will facilitate development of reduced input management systems and management strategies for long-term sustainability (see Table 111).
VIII. SUMMARY AND CONCLUSIONS Soil is a finite and dynamic living resource that acts as an interface between agriculture and the environment and is vital to global function. Soil health can be defined as the continued capacity of soil to function as a vital living system, within ecosystem and land-use boundaries, to sustain biological productivity, maintain the quality of air and water environments, and promote plant, animal, and human health. Advantages to giving value to soil health and its assessment include: (i) importance as a resource for evaluation of land-use policy, (ii) use in identification of critical landscapes or management systems, (iii) use in evaluation of practices that degrade or improve the soil resource, and (iv) utility in identifying gaps in our knowledge base and understanding of sustainable management.
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To assure the sustainability of agricultural management systems, producers and land managers must be included as active participants in the quantitative and qualitative assessment of soil health. Present research and education needs critical to assessment and enhancement of soil quality/ health include:
I . Coordinated development of standards for soil quality/health by national and local agencies and farming interest groups to assess sustainability changes with time. This requires establishment of reference guidelines and thresholds for indicators of soil qualitylhealth that enable identification of relationships between soil measures and soil function which permit valid comparisons across variations in climate, soils, land use, topography, and management systems. This will also require identification of appropriate scales of time and space for assessment of soil quality/health and development of standardized protocols for sampling, processing, and analysis. 2. Development of practical approaches and tools for on-site assessment of soil quality/health by farmers, researchers, extension, conservationists, and environmental monitors that can also be used by resource managers and policy makers to determine the sustainability of land management practices. We are beginning to realize that soil health, by its broadest definition, is inseparable from issues of sustainability. The challenge before us is to develop holistic approaches for assessing soil health that are useful to producers, specialists, and policy makers in identifying agricultural management systems that are profitable and environmentally benign, and which will sustain our soil resources for future generations.
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PEWTOREMEDIATION OF SOILS CONTAMINATED WITH ORGANIC POLLUTANTS Scott D. Cunningham,’ Todd A. Anderson,* A. Paul Schwab,’ and F. C. Hsu’ 1
DuPont, Environmental Biotechnology, Glasgow Site, Newark, Delaware 19714 ZPesticide Toxicology Laboratory, Iowa State University, Ames, Iowa 5001 1 3Kansas State University, Department of Agronomy, ’r‘hrockmorton Hail, Manhattan, Kansas 66506
I. Introduction A. “Remediation” of Soils Contaminated with Organic Pollutants B. T h e Process and Goals of Soil Remediation C. The Economics of Remediation 11. “Phytoremediation” A. Concepts and Definitions B. Apparent Advantages of Phroremediation 111. Xenobiotics in Soil A. Pollutant Distribution B. Pollutant Availability IV. Plants as Remediation Structures for Organic Pollutants A. Xenobiotir Uptake into Roots B. Xenobiotic Fate in a Plant System V. Phytoremediation ex P h t n A. Fx Plunu Enzymatic Effects B. Plant-Associated Microflora in Remediation C. Plant-Produced Physiochemical Effects VI. Modeling Phytoremediation VII. Practical Considerations A. Site Conditions and Limitations B. Agronomics C. Limitations to Phytoremediation 55
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WII. Current Phytorernediation Research and Development A. Petroleum Contamination B. Chlorinated Hydrocarbons C. Pesticides D. Biotechnological Improvements in Phytoremediation IX. Conclusions References
I. INTRODUCTION
A. “REMEDIATION” OF SOILSCONTAMINATED WITH ORGANIC POLLUTANTS The remediation of soils contaminated with organic pollutants is a global problem that consumes considerable economic resources of industries and governments alike. Contaminated soils are widespread. They are found on all continents and are often associated with centers of population, petroleum handling and storage areas, and areas of significant manufacturing activities. It is estimated that over the next 30 years 750 billion U.S. dollars will have to be spent in the United States alone to remediate contaminated sites to current legal standards (Russell et al., 1991). Although this figure contains both soil and water remediation costs, current breakdown (at 25% soils and 75% groundwater) suggests that in the United States alone $6 billion/year will be required to remediate soils over the next 3 decades. For the most part soils contaminated with organic pollutants are remediated using a diverse set of thermal, chemical, and physical methods that strip the contaminants from the soil (Nyer, 1992). In addition to these physical and chemical treatment methods, microbial-based remediations have become more common in the last 2 decades (USEPA, 1992). The first biotreatment systems targeted petroleum contaminants in shallow groundwater. These early systems relied on stimulation of naturally occurring aerobic populations to degrade the contaminants; for the most part this was accomplished by adding nutrients and increasing oxygen flux within the contaminated zone. Microbial remediation of groundwater contaminants has since expanded to target other organic pollutants and anaerobic processes as well. Concurrent with these developments in groundwater bioremediation, some soil bioremediation techniques have been developed including slurry reactors, composting, and “bioventing.” However, all biological-based processes remain a relatively minor component of the total remediation market on both a volume and a cost basis (Staff, 1993). All remediation techniques are done either in place (in situ), or by removing
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the contaminated material for treatment (ex siru). The nature of the contaminant and the soil matrix often determine the type and the location of remediation that are appropriate. Certain contaminants are inherently more easily remediated than others based on their chemical, physical, or biological properties. Contaminants that are either water soluble or volatile are often remediated by pushing or pulling air or water through the matrix. Techniques that fall into this class include sparging, vacuum extraction, and leaching. Other contaminants, however, are inherently much more difficult and expensive to remediate. These contaminants tend to be relatively nonvolatile and non-water soluble. Some engineering-based remediation techniques work to increase the extractability of the more difficult contaminants by the application of heat, surfactants, or acids, or by physical manipulation; however, intensive handling processes tend to greatly increase costs and are fraught with difficulties. All remediation techniques either remove the contaminant from the polluted matrix in a process called “decontamination” or sequester the contaminant via “stabilization.” Decontamination techniques include thermal desorption and matrix removal. Alternatively, the chemical and physical nature of the matrix may be altered to sequester the contaminant in the matrix, a process known as stabilization. The most common stabilization technique is through the addition of cement. Stabilization techniques are often acceptable in cases where reduced biological availability and off site movement can be demonstrated to the satisfaction of the regulatory and public community. All stabilization techniques are followed by site management practices that continue to contain the pollutant on site and further reduce future environmental risks. Although decontamination and stabilization techniques have different costs, and reach different endpoints, the term “remediation” is accurately applied to both. In many cases site managers would prefer decontamination processes. This type of remediation allows for increased flexibility in future land use planning and increases the value of the property. Certain sites, however, due to the nature of the pollutant, site location, extent of contamination, or the human and environmental risks involved in excavating, are potentially more appropriately remediated through a sequestration/stabilization technique than through a decontamination technique. In general, surface soils contaminated with either volatile or water-soluble organics tend to be poor targets for the development of new remediation technologies. This is because many of the existing techniques are relatively inexpensive and the extent of such soils is relatively limited. Therefore, we recommend that new remediation technologies satisfy the following criteria: ( 1 ) they must have relatively low cost; (2) they must address larger surface areas of soil; (3) they must deal with relatively immobile contaminants; (4) they should address those problems which currently have the greatest economic significance; and ( 5 ) they must meet regulatory requirements.
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With those criteria in mind, and as a background for our ongoing research programs and this chapter, we surveyed current industrial and governmental remediation expenditures, assessed future liabilities, and examined the potential for accidental releases in the future. This body of information indicates that appropriate targets for the development of new in situ, surface-soil remediation technologies for soils contaminated with organics are (in rank order of volumes to be remediated as well as expected economics): 1. 2. 3. 4.
petroleum products and by-products industry-specific chlorinated organics (PCBs, dichlorobenzenes, etc.) industry-specific nitroaromatic compounds (TNT, DNT, etc.) pesticide residues that are historic, “off label,” or accidental spills.
The development of an effective remediation system based on green plants, for these contaminants, under the above criteria would seem to be technically feasible, economically viable, and socially responsible. It should also be mentioned here that in terms of remediation opportunities, surface contamination by inorganics represents an excellent opportunity for similar new technology development. This is especially true as there are still fewer viable technologies for the remediation of inorganic contaminants of surface areas. Progress in the phytoremediation of inorganic compounds has recently been reviewed (Salt et al., 1995; Kumar et al., 1995; Baker et al., 1991) and is currently the subject of numerous laboratory and field programs.
B. THEPROCESS AND GOALS OF SOILREMEDIATION Air and water pollution is almost universally regulated in the developed world. Air and water pollutants are transported across individual property and national boundaries. There is a long-standing public mandate to regulate pollutants that affect the air and water that we share. In addition to the public mandate, there is strong legal precedent for such regulation. On this legal and popular basis many countries have passed “Clean Water” and “Clean Air” regulations. These regulations provide a solid regulatory framework from which developers of new technology can derive targets and measure effectiveness. The legal precedent for regulating soil contaminated with only sparingly soluble compounds is more ambiguous. The legal precedent in many countries is that soil is owned by the individual and not held for the “communal good.” At times, the regulation of contaminated soils has directly clashed with historic property rights. The regulation of contaminated soils is therefore more complex. In cases where the soil can be demonstrated to have a clear impact on surface or ground water, this often sets the standards for legal debate. For soils with rela-
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tively immobile contaminants the regulatory framework is more ambiguous. It is not surprising therefore that there is considerable variation in the definitions of both “contaminated’ and “clean” soils. The United States, like many countries, lacks a “Clean Soils” Act. This nonuniform legal situation places developers of new technology in an uncomfortable place with no clear idea of the “problem” definition or when “acceptable goals” or “endpoints” have been obtained. Similarly contaminated soils are treated in remarkably dissimilar manners depending on the state, the proximity to population and water Bow, and the method of contamination. A petroleum spill at a well head is regulated differently than the same material spilled in a refinery. In the first case the most liberal state’s clean-up requirement goes into effect when petroleum levels exceed 1% (10,000 mg/kg) by weight. In the most stringent case, clean-up is required when the petroleum levels exceed 10 mg/kg. These two scenarios present two vastly different remediation challenges with potentially identical contaminated matrices. Remediating soil from 50 pg/g TPH (total petroleum hydrocarbon) down to a level of 9 p.g TPHlg soil presents a very different design criteria than remediation from 4% oil down to 1%. In addition, these regulatory inconsistencies dramatically affect the total surface area to be remediated at a given site. Differences in clean-up levels also exist depending on the physical location of the site within a state. One of the greatest concerns of regulatory agencies is risk and risk management. When the contaminants have been identified for a given site, risk is often determined based upon the potential health hazard of the chemicals and the use of the land/soil. Noncarcinogenic contaminants are rated according to a hazard index which is generally based on acute toxicity. Carcinogenic compounds are ranked according to the concentrations necessary to induce cancer. The “point of departure” for the carcinogenic compounds is the concentration that causes one extra case of cancer in a population of one million assuming long-term consumption or exposure. Land use is a very important component of the determination of exposure. Soil near residential housing may be viewed in a different light than soil that will be used as industrial fill. In some states, for example, maximum allowable TPH content for residential areas is 100 mg/kg, whereas concentrations of 300 mg/kg are acceptable for industrial areas. Regulatory assessments of contaminated soil will also determine whether or not a new technology is acceptable based on the time required to complete the remediation. The overriding factor in such a decision may be whether the time required for the remediation will create an unacceptable risk. Two identical sites with different land use (an area adjacent to a day care center or an industrial park) can require different technologies based on speed of obtaining the endpoint. The developers of new technology should be well aware that the goals, constraints, and expected (or achievable) endpoints are currently highly variable. Despite the variability in regulatory guidelines between states, the remediation
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process will follow a fairly predictable chronology. The process always begins with an event, an analytical result, findings of a construction project, etc. Analytical data is collected, and if analytical results are higher than predetermined cutoffs, a report is made to the regulatory authorities. The site is then sampled more extensively and the nature, extent, and potential impacts of the contaminant are assessed. Then, in discussion with the authorities, a decision on the need for remedial action is made. At this point, the group responsible for the site further investigates the site and proposes to the regulators a suggested remediation plan. After a series of discussions is held, decisions are made on the method, timing, endpoints, and monitoring that will occur. If the clean-up occurs under a regulatory order, a legal document is produced (e.g., “ROD or record of decision) which commits the parties involved to a particular course of action. Once a ROD is signed there is considerably less flexibility in what new technologies can be tried on a given site. There are occasions, however, that a clean-up occurs under a voluntary arrangement and considerably more freedom to experiment on these sites may be possible. The developers of any new remediation techniques would be well advised to become involved in the remediation process as early as possible.
C. THE ECONOMICS OF REMEDIATION For the most part soils are remediated by engineering techniques. The nature of these techniques depends on the volume of the soil to be remediated, the physical and chemical properties of the pollutant(s), and the type of soil/sludge to be remediated. Costs vary greatly with the remediation system, ranging over three orders of magnitude. Processes that rely on in situ water flushing (“pump and treat”) or vapor stripping tend to be the most inexpensive with total project costs running at roughly $10 per ton treated. In general these processes are slower and costs are spread out over multiple years. Unfortunately, they also may only be containment techniques as many contaminants are only slowly removed from the soil in this manner. Calculations show that these systems may have to run for decades (and in some cases centuries) in order to finally remediate the contaminated area. Ex situ treatments are quicker, but are generally considerably more expensive. Cost ranges for ex situ treatment projects range from $40/ton to over $800/ton. Costs for small projects with high site management expenses and little economy of scale are the highest. Petroleum-contaminated soils are often a favorite remediation target. Costs associated with low-temperature thermal desorption of these soils run from $75 to $125/ton, and high-temperature thermal desorption ranges from $300 to $450/ton. Some of the more exotic treatments (e.g., in situ vitrification for radionuclides) can reach several thousand dollars per ton. All costs are dependent on the total volume to be remediated and can vary with
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such factors as: (I)proximity to incineration unit, landfill, and stabilizing agents, (2) state regulations, and (3) supply and demand. Like any business, the remediation business operates under free market economics, including volume discounts, sales, and “one time only if you act now” sales. Some readers might be surprised to know that landfill operators actually have sales on landfill space, while at other times shortages develop. Altogether, however, it is not uncommon in the United States to spend 1 million dollars/hectare to remediate down to a depth of half a meter. The fact that remediation is expensive is the driving force behind the search for new technology. If remediation were inexpensive, all the contaminated soil in industrial countries would obviously be remediated. An inexpensive new technology would not only save money on the remediations that do occur, but also vastly increase both the volume of soil that is remediated and the rate of soil clean-up in industrialized countries, and possibly even decrease the number of lawyers involved in “not remediating.”
11. “PHYTOREMEDIATION”
A. CONCEPTSAND DEFINITIONS The generic term “phytoremediation” consists of the Greek prefix phyto (plant) attached to the Latin root remedium (to correct or remove an evil). “Remediation” in this case encompasses all of the discussion in Section 1. Although there is some variation in the term phytoremediation as currently used, we define phytoremediation as the “use of green plants and their associated microbiota, soil amendments, and agronomic techniques to remove, contain, or render harmless environmental contaminants.” The term phytoremediation, like the term remediation itself, is rather loosely divided into processes that decontaminate the matrix (extract, degrade, volatilize, etc.) and processes that stabilize the contaminant in the soil to reduce or prevent further environmental damage (sequester, solidify, precipitate, etc.). Phytoremediation technical concepts are borrowed from many years of work by other researchers in the areas of land reclamation, landfarming of oily wastes, waste water engineering, soil chemistry, plant physiology, and agricultural pesticides. The use of plant-based (or “phyto”) remediation systems is not new. In fact the first plant based system was installed over 300 years ago in Germany for the treatment of municipal sewage (Hartman, 1975). Since that time, overland flow systems, spray irrigation systems, and constructed wetlands are common for the secondary treatment of municipal sewage waters. The concepts and principles are well understood and there are numerous companies who actively design and
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install such systems. These include reed-bed filters (Green, 1992), constructed and natural wetlands (Knight et al., 1992), and systems designed around floating plants (Buddhavarapu and Hancock, 1991). Although all of these systems are designed primarily to remove municipal waste water contaminants, there has been some work at describing the efficiency of removing industrial Contaminants as well (Wolverton and McDonald-McCaleb, 1986; Winter and Kickuth, 1989). Significant decreases in concentration of a wide variety of industrial pollutants have been shown to occur as the water passes through such vegetated beds. In recent years there has been a concerted effort to extend this concept beyond municipal water treatment systems. Using some of the same principles and plant species familiar in water treatment, the concept has been expanded to purify shallow groundwater in order to prevent off-site migration of the pollutants. There are now dozens of demonstration projects underway. Collectively, these techniques can be referred to as bio-curtains, bio-filters, or rhizo-filters. Trees, with deep roots and high transpiration rates, are being field tested to address landfill leachates, pesticide contamination, and plant nutrient elements such as N and P (Schnoor et al., 1995). Commercial names of such endeavors include EcolotreeB cap, Treemediation, and Rhizofiltration. All of these processes have targeted relatively water-soluble contaminants migrating in surface or shallow subsurface water flows. In situ, surface soil decontamination by plants without a flowing water phase is considerably less well documented, both in the lab and in the field. The use of plants in the remediation of contaminants in the air is also receiving significant attention. In one sense, we have long recognized our mammalian dependence on plants for remediation of our air. Two centuries ago, Priestley’s classic bell jar experiments containing a mouse and green plant proved that plants “remedy an evil” in our air. Followed shortly by Lavoisier’s observations, these early researchers concluded that animals give off “some sort of poison, and that the green plant renioves this poison” (Winchester, 1965). The demonstration of the phytoremediation of contaminated air (beyond CO,) is more recent. Plant leaves, with their waxy surfaces, absorb lipophilic volatile compounds including priority pollutants (Keymeulen et al., 1993) and PAHs (Simonch and Hites, 1994). They or their associated microflora have also been shown to act as biofilters for a number of indoor air contaminants. (Raloff, 1989). Over the last 2 decades a number of patents have been filed on novel designs of plant-based air biofilters intended for the home and office environment. A patent search also reveals one ambitious patent for “vegetating the external surfaces of buildings to remediate urban and industrial air pollution” (Dittmar, 1976). In addition to patent activity, a number of serious attempts at further quantifying the scientific basis of the phenomena and the extent of the effect on air contaminants are being made throughout the world. One particularly promising area of research is the genetic engineering of plants to improve the organic uptake, as well as uptake and metabolism of NO, and SO,, products of the internal combustion engine.
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The use of stationary plants to remediate streams of air and water that pass around them is relatively well advanced in comparison to the use of plants in the remediation of contaminated soil. This is partly due to the lower regulatory oversight of contaminated soils, but also reflects the fact that there are significantly more challenging dynamics involved in the mass flow, kinetic, and analytical constraints which are present in soil remediation. Within the research community concerned with the phytoremediation of contaminated soils, research on the use of plants to remediate inorganic contaminants is progressing at a more rapid pace than analogous research with organic contaminants. There are two inherent reasons for this trend: (1) Soils contaminated with inorganics have far fewer remediation alternatives than soils similarly contaminated with organics, and ( 2 ) with relatively few exceptions, inorganics are immutable. Both regulators and scientists alike can follow the fate of inorganics in a soil-plant system with relative confidence. This is usually accomplished by either direct analysis (XRF) or acid digestion and then a spectrometric analysis (ICP, ICP-MS, or AA). Organic. contaminants present both an inherent analytical challenge and an opportunity to the field of bioremediation. The analytical challenge is that following both the fate and the effect of a contaminant in a soil-microbe-plant system can be exceedingly difficult. Not only does the compound undergo physical, biological, and chemical changes in the process, but subtle changes in the matrix (plant and soil) alter chemical extraction efficiencies. Plants are themselves complex organic matrixes that can bind or mask the presence of a contaminant. ldentifying the contaminant (as well as its metabolites) in this background can be a challenging task without a radiolabeled compound. This analytical quandry increases the cost of doing research, slows progress, and prevents duplicating much of the field scale work currently occurring in inorganic research with organics. Those doing research on the phytoremediation of organics envy the ability of their inorganic counterparts to search exotic places for metal-accumulating flora. With instruments as simple as an XRF or a strip of reactive paper they can test plant tissue for metal accumulation and find a tree with sap that has Ni content in excess of 25% by weight (Jaffre et a/., 1976). Unfortunately for those involved in the phytoremediation of organics, simple analytical tools (e.g., a rapid, field bioassay for benzene degradation) do not currently exist. It would be of great benefit to organic phytoremediation to be able to search the flora of the world for variations in degradative capacity with a rapid assay. Many of those involved would love the opportunity to examine the metabolic activities of different ecosystems. One such assay may be in the development stages for a few compounds (e.g., trinitrotoluene (TNT), hexachloroethane, and triaminotoluene). Working in association with the EPA, researchers have developed a rapid bioassay for plant enzymatic activity (Wolfe et al., 1995) that can now run by high school students.
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This screening tool allows the rapid examination of hundreds of plants on a daily basis rather than the relatively few plants that can be screened in the laboratory setting on a weekly basis. As mentioned previously, phytoremediation either removes the contaminant from the matrix (decontamination) or sequesters it into the matrix (stabilization). The latter process seeks to reduce potential environmental harm by reducing the mobility and availability of the contaminant. Both processes are outlined below and illustrated in Figs. 1 and 2. Literature searches on each of the words used in the figures will generally uncover a wealth of knowledge on the biological and chemical processes involved; however, for the most part, the data are phenomenological, or limited to a single well-studied compound that is not a current target of phytoremediation for other reasons. Much of what we know about phytoremediation must be extrapolated from this literature base. Little is known about actual field rates, kinetics, lower obtainable remediation limits, and vegetation and soil management practices to accelerate the processes involved. Although the development of a remediation technique can be accomplished through trial and error (and many engineering techniques have been) phytoremediation R & D represents an excellent area in which basic and applied science have begun to work together to reach a potentially valuable technology. 1. Phytodecontamination
Phytodecontamination is a subset of phytoremediation in which the concentration of the contaminants of concern in the soil is reduced to an acceptable level through the action of plants, their associated microflora, and agronomic soil techniques. Figure 1 shows processes involved in this type of remediation. The inherent processes behind these techniques are further described below.
Phvtodecontamination Processes
Phvtovowir ation
\lCJ
\
Phvtoextractlon \ Harvest and Destruction
(Phvtod-
‘ n
K
.
Contaminant
Microbial metaboiisr
Figure 1 Naturally occurring processes involved in phytodecontamination.
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Phytoextraction. Absorption of the contaminant into the plant tissue and subsequent harvesting for destruction. Phytovolatilization. Plants or their associated microbial activity help to increase the rate of volatilization of a contaminant from the contaminated soil. The volatilization occurs from the plant shoots or roots, as well as from the soil surface. Phytodegradation. Plants take up the contaminant and metabolize it to an environmentally benign material. Rhizo(sphere)degradation. Plant roots, their associated microflora and/or excreted products destroy the contaminant in the root zone. 2. Phytostabilization
Figure 2 lists processes involved in the stabilization and sequestration of contaminants in soil. The acceptability of this type of remediation rests on processes that must be demonstrated as acceptable to the satisfaction of scientists, regulators, and the general public. These processes include the reduction in biological availability to all potential receptors as well as the off-site migration of the pollutant. In general, processes listed in Fig. 2 are well documented and occur naturally, but are hard to quantify. The acceptance of reduced bioavailability in risk assessment and remediation is increasingly well accepted for some metals (Pb, Cr, etc.). With organic contaminants, however, it is only beginning to be an important factor in regulatory discussions. No site, contaminated with organics, has been, or is being intentionally remediated by phytostabilization at this time. Sites contaminated with inorganics, however, are undergoing phytostabilization. There is growing regulatory acceptance of the concept of “relative bioavailability” of organics. The establishment of “environmentally acceptable
Phvtostabilization Processes Plants chosen for tolerance to site conditions, erosion and leachina control, and Door
Humification Contaminant
I
contaminant secluestration sequestration due to contaminant interaction cellwalllignins with increasedorganic matter soil mineral fraction upon aging.and weathering
Plant and microbial bind contaminant into soil humus
Contaminant
Figure 2 NdtUrdy occurring processes involved in phytostabilization.
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treatment endpoints” relies on this concept. The regulatory community also seems to be moving toward a risk-based framework for regulation of contaminated soils. As reduced bioavailability is a measurable and demonstratable phenomenon, the concept fits well into this emerging framework. In addition, the emerging focus on “intrinsic or natural bioremediation” invites discussions of the bioavailability concept. Ongoing research by academic, private industry, and regulatory authorities suggests that processes which involve demonstrated reduction in bioavailability to human and environmental receptors will be of increasing value. This research area is increasingly important in the determination of the fate and effect of agricultural products and environmental toxicants, and will also be important in pending legislation. Definitions are provided here for clarification. Humification. Incorporation of the contaminants into soil humus resulting in lower bioavailability. Lignification. Toxic components become irreversibly trapped in plant cell wall constituents. Irreversible binding (aging). Compounds become increasingly unavailable due to binding into soil.
B. APPARENT ADVANTAGES OF PHYTOREMEDIATION The traditional remediation community does not appreciate plants. In a technical and legal sense they fall under the category of “debris” which must be removed and treated prior to remediation. For communication purposes, with a somewhat skeptical engineering community, we have found it helpful to redefine green plants as “solar-driven pumping and filtering systems” that have “measurable loading, degrading, and fouling capacities.” Roots may be described as “exploratory, liquid-phase extractors that can find, alter, and/or translocate elements and compounds against large chemical gradients” (Cunningham and Berti, 1993). Such definitions serve as the basis for modeling efforts and economic evaluations, and provide research directions to biologists. Involvement with traditional remediation engineers is vital to the success of this new technology. Their educations and backgrounds are beneficial and may help direct research in areas not normally covered by plant biologists (e.g., supplying kinetic parameters for whole plant processes). They are also the eventual customers for any technology that phytoremediation produces. Cost effectiveness is believed to be one of the greatest apparent advantages of phytoremediation. Agronomic techniques are considerably cheaper than costs mentioned above. Farming costs are listed by county extension agents and often
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are 1 or 2 orders of magnitude less than the ex situ engineering techniques discussed in Section I. Projected phytoremediation costs also compare favorably to other bioremediation techniques. For petroleum hydrocarbons, microbial remediation techniques range from $20,000 to $60,000 per hectare (15 cm depth) in a fully contained facility. In situ microbial remediation costs (when applicable) are $7500 to $20,000 per hectare. These costs can escalate dramatically depending upon the level of monitoring (analytical costs), security (fence installation), and safety requirements (level of personal protection equipment worn). Because of reduced management and soil manipulation, phytoremediation cost estimates range from $2500 to $15,000 per hectare (15 cm depth). Low-cost ranges are for field activities associated with oily contaminants which do not require extensive safety precautions, while the upper ranges include the costs of moderate levels of analysis, consulting services, and meeting regulatory and safety standards. Difficult sites and high contaminant concentrations can increase these estimates by an order of magnitude. It is interesting to note that using some cost accounting protocols, all agronomic costs (total planting, tillage, fertilization, and harvesting cost) can be insignificant in comparison with associated costs (e.g., site management, regulatory reporting, and analytical). Relative to other forms of bioremediation, plants are an attractive benefit of phytoremediation. Phytoremediation revolves around exploiting the natural ability of environments to attempt to restore themselves. In most cases there is little public or regulatory opposition to such a line of reasoning. In addition, a field of plants provides a direct, visual bioassay that implies that the soil is relatively free of biologically damaging materials. They are also more aesthetically pleasing than any other bioremediation technique. Biofilms are notoriously unphotogenic and regular engineering remediation activities can look like lunar mining expeditions. Unlike microbial-based degradative systems, the presence and condition of plants are easily monitored. A casual glance can tell that they exist, are growing, or need water or fertilizer. The microscopic size of microbes does little to reassure an operator or concerned citizen that remediation is actually underway. This lack of monitoring capacity affects more than confidence. It also hampers the use of genetic engineering technologies in the field. Although the technical data base for transgenic microbes is much greater, the release of genetically engineered plants is now much more common than similar releases of genetically engineered microbes. Field trials with recombinant plants are now common in almost every state; over 1000 trials have been completed within the last 24 months. Engineered food products (tomatoes) have also now been introduced into the market place, with viral resistant fruits and vegetables, engineered cotton, and others not far behind. The ability to use genetic engineering tools in phytoremediation is expected to provide great benefits.
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111. XENOBIOTICS IN SOIL
A. POLLUTANT DISTRIBUTION The distribution of a pollutant in the soil is important to both the feasibility and the design of a phytoremediation system. Often identical pollutants have different distribution patterns within the soil. These patterns are based on the age, source, and nature of the spill or release as well as the type of soil. Rainfall and temperature patterns are also important in the final distribution pattern. The authors have had numerous opportunities to assess the feasibility of phytoremediation at varying sites. Petroleum overland disposal, pipeline breaks, tank ruptures, and underground tank leaks may all provide similar contaminant composition, but differing profiles present very different needs in terms of system design. In certain cases, the nature of spill events commonly associated with a particular contaminant have led to specific and unique distributions of contaminants that either preclude phytoremediation or make it an obvious choice to explore. Chlorinated solvents are one such choice. Although chlorinated solvents such as trichloroethylene (TCE) and perchloroethylene (PCE) are some of the largest remediation targets in groundwater remediations, they represent a relatively poor choice of targets for phytoremediation. Most spills occur from ruptured drums or pipes, leaking landfills, or underground pipe leaks. These solvents are more dense than water and have low surface tension. As a result they tend to enter into the soil in a relatively small area, then diffuse through the unsaturated zone (vadose zone) above the water table, then down through the underlying aquifer where they may sit as a dense, non-aqueous-phase liquid (DNAPL) on the bottom of the aquifer. The resulting contaminated zone is narrow at the top and increasing in girth as it goes deeper into the profile. The largest concern with these contaminants is aquifer contamination. As roots do not penetrate deep into even shallow aquifers to reach the DNAPL, phytoremediation possibilities are severely limited. In some rare instances a surface confining clay layer, aquitard, or hard pan may retard downward movement and make these chlorinated solvents a more attractive target; however, the total area to be remediated (consistent with our basic criteria for phytoremediation above) is low. Such generalizations concerning pollutant distribution and mode of entry should be considered prior to targeting a specific pollutant or pollutant class even if a cursory evaluation suggests that the contaminant is technically and economically viable as a target for phytoremediation.
B. POLLUTANT AVAILA~ILITY Perhaps the most important parameter in the assessment of all biological-based remediation strategies is the determination of pollutant availability. This is cur-
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rently an area of much confusion at both regulatory and scientific levels. To date, there are no clear, generally accepted protocols to determine relevant bioavailabilities for phytoremediation, microbial remediation, and even risk assessment at a site. At the scientific level it is clear that pollutants can be sequestered in soil in such a manner that they do not interact with soil biota, and cannot be metabolized, degraded, or, in some cases, even extracted with solvents. It is equally clear that there are different degrees of this sequestration and that our knowledge of the sequestration that occurs in soils and its biological relevancy is still imperfect. It is not uncommon in microbial based bioremediation treatments to encounter contaminated soils that should be (but are not) amenable to bioremediation. In some cases, despite the known degradability of the constituents, lab and field efforts cannot remediate the soil to meet regulatory guidelines. In one such case a heavy clay soil was contaminated with gasoline tank residues to a level of 10,000 p g / g TPH (total petroleum hydrocarbons as measured with soxhlet/Freon extraction and IR spectroscopy). The regulatory clean-up limit was set at 500 pg/g TPH after negotiation. A pilot project was begun with the contaminated soils. After extensive composting with numerous bulking agents, commercial innoculants, nutrients, and monitoring water and air fluxes the soil obtained a level of only 5000 k g / g TPH. A subsample of this soil was then put into shake flasks for up to 6 weeks with nutrient amendments known to induce microbial degradation of the residual hydrocarbons. The resulting TPH measurement was reduced to only 3000 p g / g , more than sixfold the negotiated limit. This apparent failure is not uncommon. There were no known microbial toxicants, the analytical profile of the petroleum contaminants appeared to be readily degradable in other systems, and the microbial plate counts and other indicators seemed to be normal for remediation. However, the soil contamination was over a decade old and its bioavailability is in question. The soil at the end of the program was not toxic to plants, earthworms or toxic to bacteria in a Microtox assay, yet still contained over 3000 kg/g TPH residue. There is no physical, technical, or microbial reasoning to suggest that phytodecontamination of such a soil would clean it up to an endpoint lower than the best that can be obtained in lab microbial degradation studies. If pollutants are sequestered against all biological interactions, they do not leach, are not available to soil invertebrates, are not available upon animal ingestion, and have no availability to dermal and lung adsorption, there remains a regulatory question that is currently under review: Should these materials be considered a risk to the environment? Considerable experimental and regulatory thought is now being given to this question. Protocols for bioavailability measurements will have to be developed, and some regulatory consensus will have to be formed. The acceptability of phytostabilization as a remediation technique
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depends directly upon the results of this consensus-building process, and upon the answers to the above regulatory questions. Most of our basic understanding of bioavailability of organics in soil comes from the development of agricultural pesticides, primarily soil-applied materials. When organic compounds are applied to soil they can be destroyed (chemically or biologically), lost from the system (leached or volatilized), or bound into the soil matrix. The processes are reasonably well understood from both an empirical and a theoretical standpoint (Sawhney and Brown, 1989). This same knowledge also forms a sound basis for phytoremediation. Soil binding occurs as charged compounds are bound to ion exchange sites and lipophilic compounds associate with soil surfaces (McBride, 1994). As most priority pollutants are uncharged at neutral pH, the greatest effect is the interaction of lipophilic compounds with solid surfaces and organic matter in the soil. In most cases, soil organic matter provides the most important absorptive capacity in binding a wide range of relatively lipophilic compounds (Chiou, 1990). An acknowledgment of this effect is written into many agricultural products’ ‘‘labels’’ (notes provided by the manufacturer on the use of the product). In high-organicmatter soils the suggested application rate is greatly increased. This is due to the high binding and reduced bioavailability of the product in the presence of organic matter. This bioavailability is reduced for plant, invertebrate, and microbial systems (the three targets of agricultural products for which there is ample evidence) as well as leaching tests. As a result, all forms of biological decontamination, including phytodecontamination, would be predicted to be less efficacious in soils with high organic matter contents; conversely, phytostabilization should be more effective and applicable. In addition to the effect of organic matter on bioavailability of organic pollutants in soil, the amount of clay in a soil and clay type are also known to have an effect on adsorption and availability of the compound (Chiou, 1990; McBride, 1994). Clay particles have large surface areas so that high-clay soils have larger surface areas onto which the organic may adsorb. In addition, certain types of clays have the potential for interlayer entrapment of organics as well. All changes in soil texture that decrease average particle size generally slow down diffusion of a pollutant to a root by increasing the tortuosity of the path of water flow, and by decreasing diffusivity. Soil structural characteristics (e.g., soil aggregates) may also help lower the biological availability of the organic contaminants. In some cases the pH of a soil may also affect the bioavailability of compounds, especially those that are charged and whose sorption is dependent on ionic binding sites. The effect of soil constituents on binding and availability of a compound has been well studied in recent years. One additional variable that has received less attention and is exceedingly important in discussing hazardous wastes is the effect of aging on bioavailability. In addition to the types of binding that occur in
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shorter time frames, the effect of long-term aging must be considered. Many of the sites we currently seek to remediate are decades old. With better preventive practices we are creating fewer new ones, hence our site focus has been to “well aged residues” in soil. It is becoming increasingly clear that in some, if not all, cases, bioavailability and remediation potential decrease over time (Hatzinger and Alexander, 1995). As this laboratory study indicates, this decrease in bioavailability can come without any apparent decrease in the pollutant concentration as determined by regulatory extraction protocols. The exact nature of this sequestration effect has yet to be determined. This is an extremely active area of research in a number of labs. Despite this lack of complete knowledge over the cause and nature of the aging effect, it clearly can be demonstrated that desorption kinetics and leachability (Pignatello, 1989), as well as microbial availability (Steinberg et a l . , 1987; Hatzinger and Alexander, 1995), decrease with time. Better protocols to quantify this change in availability are needed. Current analytical methods by which pollutants are detected and regulated present a fundamental problem to all biological based remediation methods. In actuality biological and chemical “availability” is a continuum from “freely available in the water phase” to “irreversibly bound.” All extraction protocols remove varying degrees of “available” material. Unfortunately for those involved in bioremediation, regulatory protocols developed in the late 1960s and early 1970s codified analytical procedures that measure neither the total amount of contaminant present nor that fraction which is biologically relevant. A reexamination of analytical protocols to more accurately reflect risk, relative bioavailability, and remediation potential and need would greatly aid all forms of bioremediation. Regulatory analytical protocols that cannot differentiate the bioavailable fraction from that which is tightly sequestered form an inherent bias against all forms of bioremediation, including phytoremediation. A stronger case could be made for keeping the current regulatory extraction protocols (e. g. soxhlet with boiling Freon) if it cculd be demonstrated that they remove all the contaminant present. They do not. These protocols generally give near quantitative recoveries with recently spiked environmental samples that are low in organic matter and have low biological activity. Recoveries of spiked materials in these cases are near 100%.However, with more normal environmental samples that contain higher organic matter, and are biologically active, aged soils’ analytical recoveries drop dramatically. Newer techniques, such as supercritical fluid extraction (SFE) and analytical high-temperature thermal desorption (HTTD), tend to remove a greater fraction of the pollutant in these cases. The absence of total recovery is more a technical observation than a source of potential risk, since these tightly bound materials seem to have little biological relevancy. Pollutant residues that are not extractable by solvent extraction, yet are extractable by more advanced techniques, are referred to as “irreversibly bound,”
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“unavailable,” or “bound residues” (Pignatello, 1989). Materials such as these are currently unregulated, and seem to represent materials with low bioavailability and little risk to the environment (Alexander, 1994).
IV. PLANTS AS REMEDIATION STRUCTURE FOR ORGANIC POLLUTANTS
A XENOBIOTIC UPTAKE INTO ROOTS Being mostly sessile for their entire life, higher plants have to cope with a continuous onslaught of many biologically active materials coming in contact with the root system. As with naturally occurring organic compounds, organic contaminants are taken up by plant roots, then with varying degrees are sequestered, metabolized, and translocated to shoots. Although phytoremediation of organic contaminants is a relatively new field, studies have been conducted on the interaction of organic pesticides with plants for decades. Pesticide studies are particularly instructive in that many have been conducted with closely related analogs in a single chemical class, with systematic chemical variations, and often under carefully controlled experimental conditions. The nature of the compounds, the attempts at elucidating basic principles, and the economic importance led researchers in this field to come to some general conclusions about the behavior of organic chemicals in plant and soil systems. Soil has three phases: solid, liquid, and vapor. Plants obtain contaminants from the soil most readily from the liquid phase, but in some cases vapor phase uptake is important. The evidence that plants can take up organic chemicals without a water- or gas-phase intermediate is less convincing; however, precedent exists for convincing uptake in this manner for other systems.
1. Vapor Phase Vapor-phase movement of organics into plants from the surrounding soil can be an important uptake mechanism for some compounds, even those with relatively low vapor pressure (Mayer ct a l . , 1974). Some soil insecticides rely on this delivery mechanism. They are placed into the soil and slowly volatilize as a crucial means to get to pests that attack roots, such as the corn root worm (Diubrorica sp.). Once in the gas phase, the organics may be absorbed by the roots or leaves. The equilibrium between the gas and liquid phases for a compound can be described by Henry’s Law,
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where K , is the Henry’s Law constant, [A(aq)J is the concentration of the component in the solution phase, and PA is the partial pressure of A in the gas phase. Ryan et ul. (1988) compiled a list of Henry’s constants for priority pollutants. Some attempts have been made to define a value for the Henry’s constant above which volatilization is an important mechanism for bioavailability, but one must consider other chemical properties such as adsorption to the soil, microbial decomposition, and solubility in water. For example, the polyaromatic hydrocarbons (PAHs) are generally considered semivolatile. However, the microbial decomposition of the lower molecular weight compounds (such as naphthalene) and/or strong adsorption to soil components for higher molecular weight PAHs (such as pyrene and benzralanthracene) make volatilization from soil a minor pathway for this class of compounds.
2. Water Phase Lipophilicity is the most important property of a chemical in determining its movement into and within a plant. Lipophilicity is the “balance between the affinity of the chemical for aqueous phases and that for lipid-like phases” (Bromilow and Chamberlain, 1995). This property determines the ease of movement across plant membranes. Lipophilicity is related to the partition coefficient of the pollutant between I-octanol and water (Kow).The K , , is one of the most widely available experimental parameters for xenobiotics. The values of K O , cover the wide range from 10-2 to 10’0. For ease of discussion these numbers are often reported in log form. A small KO, is indicative of high water solubility and low lipophilicity, and high values are associated with compounds that have high lipophilicity and low water solubility. Organic contaminants in soil can transfer to roots in the soil-pore water by diffusion and by mass flow. In either case, the transfer is mediated by the aqueous phase. The spot-to-spot variations of soil texture, organic matter composition, water conditions, and distribution of organic contaminants make the root exposure to soil organic contaminants a highly heterogeneous and constantly changing phenomenon. This explains why many experiments have been conducted in hydroponic cultures in which an idealized condition can be easily created and maintained. For a hydroponic system, the concentration of the test compound can be kept constant to all parts of the root during the entire experimental period. Compounds that are most water soluble will be most available for mass transport and diffusion into the rhizosphere. However, very soluble compounds with little afYinity for the soil solids also will be subject to leaching out of the root zone and can become physically unavailable for root uptake. Marginally soluble components will not move with the water and require the root to grow near them before uptake is possible. Positively charged organics will tend to be retained by the soil’s cation exchange sites and their availability for assimilation will be reduced. Negatively charged compounds are excluded from the generally nega-
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tively charged soil surfaces, making them more available to the roots but further subject to leaching. Research on the plant uptake of organic chemicals from soil has centered on pesticides and a few high-profile contaminants (such as PCBs and dioxin). Pesticides have been emphasized because of the importance of plant uptake in their function and their potential for effects on the food chain and nontarget crops. Also, knowledge of a pesticide's environmental fate is necessary for its registration and sale. Plant uptake models have been developed using experimental data for those compounds that have been studied and these models allow us to predict plant uptake of organic compounds for which no experimental data exist.
3. Solid Phase Conceivably, an organic compound adsorbed to a soil particle can directly transfer into a plant root via solid-to-solid partitioning. This route of transfer may allow us to rationalize some putative high root uptakes of very lipophilic compounds with high vapor pressures as well (McMullin, 1993). This concept is technically feasible and well demonstrated in dermal sorption studies in mammalian systems; however, due to the vast surface area in a soil and the relatively low surface area of a root in comparison, it is hard to imagine this as a viable remediation strategy. Nevertheless, we do not wish to rule this out as either a mechanism of plant uptake or a potential remediation technology in some creative form.
B. XENOBIOTIC FATEIN A PLANTSYSTEM 1. Transport within the Plant Physiological factors of plant roots that control uptake of organic chemicals have been summarized by McFarlane (1995). Briefly, water and dissolved constituents can move easily and relatively unimpeded from the soil solution into the root's apparent free space, that area of the root outside the endodermis and the Casparian strip. The apparent free space is characterized by cortex cells with porous cell walls and many voids. The Casparian strip in the endodermis is a waxy barrier that inhibits movement of water into the interior of the root. At this point, all water, solutes, and non-aqueous phase liquids must pass through cell membrane at least two times. For most xenobiotics (both ionic and nonionic substances), this movement a p p r s to be a passive process depending upon retention by the membrane, solubility in water, and diffusion. Many observations on xenobiotic root uptake and root-to-shoot translocation have been made (Devine and Vanden Borden, 1991). However, most were aimed at describing the
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behavior of specific compounds of interest. Here we only review those studies that are instrumental in revealing the underlying general principles governing the processes. Movement of nonionized organic chemicals into the root consists of the equilibration of the solution phase outside the root with the solution in the apparent free space and sorption of the compounds onto the root surface. Shone and Wood (1974) examined the uptake of the herbicide sirnazine by barley growing in solution culture. One of the more interesting observations of their research was that the xylem translocation of simazine apparently was restricted because the concentration of simazine in the transpiration stream was less than the concentration in solution. A transpiration stream concentration factor (TSCF) was developed to describe this behavior. TSCF
=
concentration in xylem sap concentration in external solution
In the case of simazine or any other compound whose xylem translocation is apparently restricted, TSCF is less than 1. 0. Other triazines also were found to have TSCF less than 1.0 (Shone el a!., 1974). Shone and Wood (1974) also defined a root concentration factor (RCF). RCF
=
root concentration external solution concentration
This line of study was extended by Briggs et al. (1982) to a series of other pesticides (0-methylcarbamoyloximes and substituted phenylureas). They found no relationship between RCF and TSCF and suggested that root accumulation of one class (triazines) was mostly physical adsorption to the surfaces of the roots. For other classes (carbamoxyloxime and phenylureas), an empirical model was made to relate root uptake by intact barley plants to lipophilicity: log (RCF - 0.82)
=
0.77 log KO, - 1.52.
This equation is expected to vary somewhat among plant species depending upon the composition of lipids in the roots. They found that the TSCFs were less than 1 .O for all 18 compounds studied, and that TSCFs showed a bell-shaped relationship with log K , , values. TSCF
=
0.784 exp
[(log KO, - 1.78)]* -2.44
The maximum TSCF was observed for log KO, of about 1. 8. They also plotted those literature data with sufficient details and found that despite the diversity of plant species, compounds, and experimental techniques, they largely conform to the same relationships they found between log K , , and RCF and TSCF. The general conclusions of Briggs et al. (1982) were further validated by Hsu et al.
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(1990), albeit with a different maxima, on a series of herbicide (cinmethylin) analogs using a root pressure chamber technique. While the linear relationship between RCF and log KO, is immediately intuitive, the log KO, and TSCF relationship requires explanation. It seems to be intimately related to the pathways from root surface to xylem vessels situated in the stele. The root pathways for water and solutes have received reviews in the past (Weatherley, 1975). The salient feature of the main pathway is that a molecule has to cross cell membranes at least twice, once before the Casparian strip (CS) barrier on the endodermis, and once after that, to gain access to the xylem in the stele (Weatherley, 1975; Clarkson and Robards, 1975). This is referred to as the symplastic pathway. The apoplastic pathway, available only near the tip of the root where the CS has not yet developed well, usually accounts for an insignificant amount of the total flux unless the root is disturbed (Moon et af., 1986) or stressed (Hanson et al., 1985; Skinner and Radin, 1994). Compounds with low log KO, values can move through the intercellular space along with the mass flow of water until reaching the CS barrier. There, the two membrane crossing steps are slow due to low lipophilicity. On the other hand, delivery efficiency of compounds with higher log KO, values through the pre-CS barrier parts is low due to low water solubility and losses to lipophilic tissue constituents. But the rate of their membrane crossing is more rapid, compensating for the earlier slower passage. Once compounds are partitioned into lipophilic membranes interior to the CS barrier, they need to be desorbed into aqueous solution in order to go into xylem vessels. Conceptually, the entire root pathway may be simplified as a step of partitioning into the post-CS membrane and a desorption step off this membrane. The first step of aqueous-to-lipid partitioning obviously favors more lipophilic compounds. However, the second lipid-to-aqueous desorption step favors more hydrophilic compounds. The interplay of these two crucial rate-limiting steps would produce the observed relationship between log KO, and TSCF (Briggs et al., 1982; Hsu et al., 1990). In soil the idealized hydroponic condition is compromised. Sorption of organics to soils can limit their availability to roots. For many organic compounds, one can assume that the adsorption is a linear function of concentration and organic carbon content of the soil, 9e = K d f ,
c,
wheref,, is the fraction of organic carbon in the soil, C , is the concentration of the organic compound in the soil solution, and Kd is the linear adsorption coefficient. This equation suggests that as organic carbon content increases, adsorption increases and availability for root uptake decreases. It is estimated that in soil the log KO, for maximum TSCF shifts down by about two units (Hsu et al., 1990). In the field of phytoremediation, not much attention has yet been paid to the mechanistic aspects of root uptake and xylem translocation. Most studies use
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endpoint analyses of shoot contents of radiolabeled test compounds. Such studies are always complicated by shoot metabolism and severely limited by the availability of radiolabeled compounds. Simple techniques are now available to allow for mechanistic experimentation with nonradiolabeled compounds free of shoot metabolism complications. This kind of mechanistic approach can yield crucial information to help design the most optimal phytoextraction scheme for important soil organic contaminants. Conceptually, the total amount of organic compound phytoextracted into the (easily harvestable) above ground fraction can be defined as: Amount delivered to shoot = Conc. in sap
X
Volume of sap/time unit.
The xylem sap can be obtained from the decapitated stem near the soil line either by applying pressure to the root of a potted plant in a modified pressure chamber as that used in Hsu et al. (1990) or by applying negative pressure to the cut stem (Gil de Carrasco et al., 1994; Ambler et al., 1992). Two cautions need to be observed to obtain xylem saps matching those of intact plants: ( I ) adjust the sap expression rate to match that of the sap volume flow in intact plants, and (2) avoid the initial sap sample which may contain artifacts associated with a given technique (Else et al., 1994). Due to the generally simple composition of xylem saps, the presence of a target organic compound and its metabolites can be readily quantified by analytical techniques. Here the analytical challenge mentioned earlier (Section 1I.A) is quite easily met. This is evidenced by the successful quantification of low concentrations of the natural plant growth regulators abscisic acid and zeatin in xylem sap (Ambler et al., 1992; Davies and Zhang, 1991). The whole-plant xylem sap flux rate can be measured with a stem-flow gauge. The accuracy of the xylem mass flow measured with this method has been validated in many studies (Baker and Van Bavel, 1987; Steinberg et al., 1989; Dugas, 1990; Devitt et ul., 1993). The adaptability of the xylem sap expression techniques and the xylem mass flow measurement technique makes them particularly suitable for phytoextraction research. Both techniques can be used for both laboratory and field experiments with nonradiolabeled compounds. For laboratory studies, different plants can be grown in soil with the target organic contaminant. The sap flow can be measured, and xylem sap obtained for analytical quantifications. Different xylem sap concentration values from different plants multiplied by their respective total sap volume flow rates will generate the total amounts extracted by test species. The study can be run at different times of day to cover diurnal variations of xylem flow rates and sap contaminant concentrations. The method can also detect whether the build-up of the target compound in shoot causes any self-limitation of further phytoextraction by an inhibition of sap flow rate. By using these monitoring techniques, different innovative methods can be tested to see if they produce any improvements on net phytoextraction. For field studies, sap analysis and sap flow measurements can be made to existing plants in a contaminated site.
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These measurements can quickly lead to the identification of the most advantageous plant species for phytoremediation. Although this chapter deals specifically with organic compounds, the same xylem sap analysis and sap flow measurement can be effectively applied to the mechanistic study of phytoextraction of other soil contaminants, particularly inorganics. Knowing their xylem sap contaminant content as well as the chemical identity of the chelating compound(s) would greatly help us in pinpointing the potential rate-limiting step in phytoextraction. Despite these general rules, there are distinct variations among plants in their ability to accumulate organics into the roots. Selecting plant species and varieties for maximizing this trait, however, has not been approached systematically. The subject has been studied with certain pesticides, but few data exist for other organic pollutants. The best-studied cases seem to involve chlorinated organic insecticides. For example, beets, turnips, potatoes, and radishes tend to accumulate less than do carrots (Lichtenstein and Schulz, 1965). Sugar beet roots accumulated more dieldrin than carrots, potatoes, corn, and alfalfa (Harris and Sans, 1967). Different varieties of carrot can have four-fold differences in endrin uptake (Hermanson et af., 1970). More detailed comments on the influence of plant properties on root uptake of organic compounds are provided by Bell (1992) and Shimp and co-workers (1993). In addition to root uptake, the general microbial stimulation in the root zone suggests that the most logical starting point in selecting species is to focus on the plant root. Species with extensive and fine root systems should have the greatest potential for enhancing bioremediation. These roots and their associated microflora would be more apt to have greater soil/root surface contact and be able to penetrate small pores than species with a coarse taproot system. Mycorrhizae may provide additional advantages because of their fine architecture ability to increase the effective surface area of the root as well as their microbial metabolic traits.
2. Metabolism within the Plant Humankind was not the first to create biologically disruptive organic compounds and place them in a soil-plant environment. Plants and their associated microflora evolved in an environment where they were continually assaulted with a wide array of microbial and plant toxins. Fungal and bacterial toxins are well known (TeBeest, 1991; Yoder, 1980; Rice, 1974). Certain plants produce “allelopathic” chemicals that suppress the growth of other plants around them (Putnam, 1985; Durbin, 1981). In addition to plant-produced herbicides, plants also manufacture a wide range of compounds with adverse pharmacological effects in herbivores. Rotenone and pyrethroids are plant-produced insecticides. Coumesterol can alter the mammalian estrus cycles and decrease birth rates.
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Taxol can alter mammalian cell cycle (and cure some forms of cancer). The genetic induction, enzymatic metabolism, and biological effects of some of these plant components (e.g., flavonoids, isoflavonoids, and coumarins) are well studied (Stafford and Ibrahim, 1992; Cody e t a / . , 1986). Many of the initial steps in their production from the PAL (phenyl alanine ammonium lyase) have been cloned and successfully expressed in other plant tissues. This is an active area of research and other important enzyme classes, pathways, and genes remain, no doubt, to be discovered. In many cases, there is a remarkable structural and chemical similarity between a toxic xenobiotic pollutant and either a natural toxin or a specific natural enzyme substrate. This is not coincidental. Many acutely toxic xenobiotics gain their “toxic” classification because they interfere with cell processes in ways similar to natural products. Some agricultural products are specifically modeled after a natural analog (e.g., bacterial glutamine synthase inhibitors are equivalent to a commercial herbicide). It is also not uncommon to seek out plant extracts and test them for bioactive compounds in the discovery of antimicrobial, antiinsect, and phytotoxic products. All of these natural compounds are synthesized by biological systems, are tolerated by at least some members of the biological community, and are finally degraded by organisms in the environment. It is therefore not surprising to plant biochemists, pharmacologists, and traditional medicine men that plants and their associated root zones have developed significant capacities to metabolize both natural and xenobiotic toxins. Most of our knowledge about plant-based metabolism of xenobiotics comes from the development of agricultural pesticides. This technical basis dramatically skews our knowledge base for two reasons. The first is that most studies have been carried out on pesticides and not on industrial pollutants. Studies on the uptake, translocation, tolerance or metabolism of industrial pollutants represent only 3% of the literature base (Nellessen and Fletcher, 1993). Second, pesticides are generally tested on crop plants, most of which have little chance of survival in truly contaminated soils. Tabulated data again show that 77% of the studies were conducted on crop plants, and that the metabolic capacity of most weeds is unstudied (Nellessen and Fletcher, 1993). Many people unfamiliar with contaminated soils expect that such sites would be barren of all vegetation. In some cases this is true, however, on most sites hardy, tolerant, weed species exist. These “volunteers” spread out over time to establish a general cover at most sites. Often sites that are heavily polluted are colonized from the edges inward, with the rate of colonization seemingly dependent on contaminant load, physical soil factors, and general cultural conditions. Many of these sites spontaneously revegetate, the most common exception being those sites with active weed control programs. This spontaneous revegetation phenomenon is probably not unfamiliar to those who have spilled oil or gasoline on the lawn. In cases where spontaneous revegetation does not occur, fertilizer,
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loosening up the soil, and water may make dramatic improvements. We have surveyed many contaminated areas and seen this spontaneous revegetation occurring across many soil types and climates. It is also interesting to note that the species which seem to come into these areas are nearly unanimously the weed species coincident with that region, These hardy weed species tend to have windor animal-borne seed distribution techniques and can be found growing in many of the nutrient-poor, disturbed areas (e.g., road cuts, abandoned fields) throughout the region. Perhaps not so coincidentally, it is many of these species that are problematic weeds in farmers’ fields and are therefore specifically targeted in the development of new herbicides. In one author’s experience (S.D.C.) at least half of the top 10 weeds targeted for new herbicides development can readily be found as volunteers on contaminated soils in that region. As these casual field observations might indicate, and as the best efforts of hundreds of the world’s pesticide chemists can attest to, certain weed species are difficult to kill and are relatively insensitive to chemicals that easily kill crop plants. Given the skewed data base on metabolic capacity of plants, it appears obvious that phytoremediation would benefit from additional research in the study of weed metabolic capacity. The results of agricultural product research also suggest there are wide differences in the ability of plants to metabolize xenobiotics. The backbone of the multibillion dollar selective herbicide business is based on this differential metabolism. Nearly all modem selective herbicides are selective due to plant metabolism. The tolerant plant selectively metabolizes the herbicide to a nontoxic compound and remains unaffected, while the nontolerant weed species either cannot metabolize the compound or metabolizes a nontoxic compound into a toxic one, thereby committing suicide (Hathway, 1989). There are significant differences between monocots and dicots in this capacity as well as between individual genera and species that may be exploited in phytoremediation. Differences in plants’ abilities to metabolize environmental pollutants are also increasingly evident from screening at both the whole plant level (Schnoor er al., 1995; Hughes and Saunders, 1995) and the cell culture level (Groeger and Fletcher, 1988). Plant xenobiotic metabolism is remarkably similar to the types of xenobiotic metabolism that occur in mammalian livers. Relatively lipophilic materials undergo an enzymatic attack, which results in more water-soluble compounds. In mammalian systems the final disposition is often through excretory routes which plants lack. In principle, however, both plant and liver metabolic systems can be divided into the same three phases: transformation, conjugation, and final disposal. The final stage in plant metabolism consists of transport and compartmentalization of the metabolized products into cellular vacuoles, intracellular spaces, or various cell wall components (Sandermann, 1992). Contributing to the comparison is the fact that two major enzyme systems responsible for liver detoxification processes are also found in plants: ( 1 ) cytochrome P450 oxygenases, and (2)
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glutathione S-transferases. It is on the basis of these comparisons that plants have been dubbed “green livers” (Sandermann el al., 1977). Some genes necessary for the degradative capacities in plants are constitutively expressed. Many herbicides are metabolized in some plant species almost immediately and there is no detectable lag phase. For phytoremediation processes that would rely on these processes functioning, concentration thresholds would be expected to be fairly broad. In other cases, enzymatic activity is not constitutive but induced. In these cases a low level of a pollutant may have one fate in a plant and a higher level might have another. Much like certain microbial systems, the presence of one toxicant (e.g., toluene) can induce metabolic pathways that will then also degrade other pollutants. In plant metabolism, the clearest example of a parallel phenomena is in the development of compounds called “herbicide safeners.” These chemicals are applied either prior to the application of the pesticide or coincident with its application. Their purpose is to trigger the production or activation of degradative enzymes which degrade the herbicide before it has a lethal effect on the plant. Most herbicides kill plants by interfering with a specific metabolic pathway or process. Sublethal quantities of herbicide or analogs may also trigger this inducible metabolic activities in many cases. Exogenously applied inducers of metabolic activity are an underutilized laboratory and field tool in phytoremediation. A review of the area of these metabolic enhancers and other potential manipulations of plant degradative capacity is provided by Hatzios and Hoagland (1989). It is widely speculated that phytodecontamination systems might be most appropriately managed by maximizing the various stress conditions on the plant (chemical, fertilizer, planting densities, etc.). This is an approach considerably different than managing a crop for conventional yield purposes. Much more research in this area is needed prior to making informed field decisions. Despite all the above discussion of plant degradative capacities, plant metabolic systems pale in comparison to their microbial analogs. Plant systems do not have as broad a substrate range, nor can they act over as wide a concentration range, as their microbial counterparts. This can easily be illustrated by comparing the respective abilities of plants and microbes to break aromatic and aliphatic C-Cl bonds. In plants, there are perhaps four well-documented C-Cl bond breakage reactions (Hathway, 1989). In microbes there are perhaps a dozen (Neilson. 1990; Chaudry and Chapalamadugu, 1991). This type of comparison has prompted the current research to exploit plant-microbial associations and/or engineer plants for better metabolic activities.
3. Sequestration within the Plant A xenobiotic compound entering the plant need not necessarily be metabolized for successful phytoremediation. In some cases, it may be possible to use root crops with high lipid contents to absorb lipophilic organics from the soil. The
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roots would then be harvested and processed. This type of phytoextraction of certain contaminants from low-organic-matter soils may be possible. One replicated pilot scale experiment used carrots to remediate a DDT-contaminated soil. Five carrot varieties were grown, harvested, solar-dried, and incinerated. DDT levels were reported to decrease in a carrot variety dependent manner by 30 to 86% over the controls (McMullin, 1993). The need for harvesting and postharvest processing has economic consequences on the phytodecontamination scheme; however, it would appear that phytoextraction and root harvesting may be potentially viable in certain cases. There remain, however, significant questions concerning the general utility of such an approach across soil types and contaminants. Beyond phytoextraction, if plant roots could be demonstrated to take up a pollutant and sequester it into an unavailable fraction, such a process might also be useful as a basis for phytostabilization. In such a case it would be imperative that the compound be so tightly sequestered that it would be essentially unavailable even to animals that might feed on the root tissue. This process has been clearly demonstrated in the case of certain I4C-labeled pesticides that become irreversibly bound into plant roots. Such residues are resistant to exhaustive solvent, acid, alkali, and enzymatic extraction protocols. Furthermore, direct feeding to rats does not result in release of the compound into the animal’s system and the rat passes the labeled compound through the digestive system with the other nondigestible fraction of the food (Kahn, 1982). Residues deposited in the lignin fraction of the plant seem to be relatively biologically inert. As they do not appear in regulatory extraction protocols they are often considered “degraded.” Additional research is needed to determine the extent to which such materials are released upon the death and decay of the root.
V. PHYTOREMEDIATION EX PLANTA A. Ex PLANTAENZYMATIC EFFECTS As has been previously stated, plant enzymes can metabolize a wide variety of xenobiotic pollutants. Plant degradative enzymes are not limited, however, to functioning only internal to the plant root, stems, and leaves. These enzymes can also be found in their active forms both in soil and in sediments. The best characterized of these plant enzymes that occur external to the plant root are certain oxido-reductases and laccases. More. recently, however, a greater variety of active plant enzymes has been discovered in sediments far from their plant source. These enzymes include dehalogenases, nitroreductases, and nitrilases (Schnoor et a l . , 1995).
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Some of the best studied oxido-reducatases are the peroxidases. These enzymes have been identified on the external root surfaces of such diverse plants as cotton (Mueller and Beckman, 1978), wheat (Smith and O’Brien, 1979), cress (Zaar, 1979), and tomato and water hyacinth (Adler et al., 1994). In tests with water-borne contaminants the capacity of these plant enzymes to interact with phenolic, aniline, and certain other aromatic contaminants is well documented (Adler et al., 1994; Dec and Bollag, 1994). Due to analytical difficulties, similar reactions are more difficult to clearly demonstrate in soils. Initial research in this area by a number of labs, however, seems promising, and additional work is ongoing. The result of the action of oxido-reductases on pollutants is often the polymerization of the pollutant either onto the root surface or into the soil humic fraction. Contaminants bound in such a manner are no longer available to most, if not all, biological processes and ordinary chemical extraction protocols. Of particular note is the fact that they are not extracted by regulatory mandated extraction protocols. These polymeric complexes are often referred to as “bound residues.” The overall enzymatic incorporation into the polymeric humic fraction of soils has also been referred to as the “humification” process, and is depicted in Fig. 2 as part of phytostabilization. Some researchers in this field would suggest that the humification process should be listed under the phytodecontamination category, as the regulatory analytical results suggest. We consider humification a stabilization process, as certain analytical techniques, including some forms of hightemperature thermal extraction and super critical fluid extraction, have been shown to release some of these bound residues. There are significant differences between plants at the level of enzyme production. There may also be differences in the release rate and timing (age, season, stress induced, etc.) although this remains to be tested. Certain plants (e.g., horseradish, Armoruciu rusticana) are cultivated for their root enzymatic capacities. Their value as a condiment and in commercial enzyme production is derived from their peroxidase production. Their potential use in soil remediation is untested; however, they have shown intriguing possibilities in water decontamination processes (Dec and Bollag, 1994). These investigators are currently conducting a survey of plant laccase and peroxidase and initial results show significant variations between plant species. In addition to humification by enzymes directly derived from the growing plant, fungal symbionts, parasites, and saprophytes, living in conjunction with the plant and its detritus, produce a wide variety of enzymes which may be involved in humification (Bollag et a l., 1995). Individually. same of these fungal species (e.g., white-rot fungi) are specifically targeted to pollutants (Yadav and Reddy 1993) and are currently being used in some field scale remediations. Degradation capacities depend not only on the production of these enzymes,
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but also on the activity of the enzyme, their release rates, the nature of the soil matrix, and the concentration of contaminant it encounters relative to other potential substrates. It may be possible to select or engineer plants and subsequently manage their culture and agronomics to exploit the humification process. Plant selection techniques include enzyme assays, altered rooting structures (a “hairy root” phenotype by Agrobacterium rhizogenes transformation should increase soil/root surface contact), or even mechanically breaking the root tissue to release the enzymes. Additions of cofactors, catalysts, and electron sources or sinks are also both agronomically and economically possible. Many of these root peroxidase activities are greatly increased in the presence of hydrogen peroxide. One of the interesting possibilities inherent in phytoremediation research is that remediation strategies can be developed to take advantage of the expertise required in remediation. Although hard to imagine in an agronomic sense, a strategy has been proposed to: (1) cultivate horseradish, (2) rototill the crop on an annual basis (it will resprout), and (3) land-apply hydrogen or calcium peroxide when rototilling (Flanders et. al., 1995). The engineering community finds the cultivation of horseradish in hazardous waste amusing. The agronomic community is skeptical of the application of hydrogen peroxide. The strength again can be seen in a multidisciplinary approach to phytoremediation. A note for the agronomic readership: hydrogen peroxide has been a common additive to wells in contaminated aquifers to increase dissolved oxygen concentrations and speed aerobic bioremediation. A note for the engineering community: horseradish is easy to grow, invasive, and hard to kill once established. A note for all: “horserado-remediation” is not yet available for fullscale (or any other) field testing. Traditionally, all soil scientists are taught that enzymes placed into soils are quickly deactivated by clay adsorption and microbial degradation (“Unprotected protein has a short life in soil”). It would appear, however, that in some cases, these plant degradative enzymes are remarkably robust. In a unique approach to searching for the cause of the natural attenuation of many xenobiotics that is observed in the environment, an EPA sponsored team of chemists, biochemists, and engineers searched freshwater sediments for five specific enzymatic capacities (dehalogenase, nitroreductase, peroxidase, laccase, and nitrilase). Using a unique protein isolation/purification scheme and following this with the use of biochemical, molecular, and immunological detective work, this team determined that “in all cases where we found significant natural activity in the sediments we isolated plant enzymes as the causative agent” (Schnoor et al., 1995). This remarkable statement suggests that plant enzymes might have significant catalytic effects well beyond the confines of the plant itself and bolsters the general case for following phytoremediation effects well past the death of a single plant.
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B. PLANT-ASSOCIATED MICROFLORA IN REMEDIATION As mentioned previously, perhaps the most intriguing possibility in phytodegradation is the exploitation of the plant’s photosynthetic, Huid transport, and soil exploratory capacities, in conjunction with the increased degradative spectrum and capacity of its microflora. In designing systems based on plant-microbial interaction, the complexity is greatly enhanced to plant-based systems alone. What had been a tripartite question (contaminant-soil-plant) is now a question involving the four-way interactions between contaminant-soil-microbe-plant. There is a long technical history in research and development of plantmicrobial associations. Research on nitrogen-fixing bacteria (Rhizobiaceae, Anabanea, Azospirilfum), plant-growth-promoting rhizobacteria (PGPRprimarily Pseudomonus), biological disease control (fungal and bacterial), rootfeeding insect control (bacterial, fungal, and viral), and root nematode control (fungal and bacterial) has convinced many researchers that the system is too complex to control effectively. The most common experience is that soil and plant innoculants are out-competed by native microflora. This is generally true even in some symbiotic relationships (e.g., soybean and their Bradyrhizobiu symbiont) which would be expected to be easily controlled. In most cases the presence of native microHora overwhelms attempts at controlling or “biasing” microbial-plant interactions for any specific effect that we have tried to date. Undaunted, we believe that biasing rhizospheres in phytoremediation is inherently different. First, these soils are significantly compromised and have microflora that are lower both in diversity and in number than found in agronomic soils. Second, biasing rhizospheres in an agronomic sense must be done within agricultural economic scales (less than !$100/hectare).The economic limitation on hazardous waste sites can be conservatively estimated to be I to 3 orders of magnitude higher. This opens the opportunity for more intensive and expensive techniques (including drip irrigation of a continual innoculum) to be employed. Third, the molecular tools and understandings are continually improving and now allow for additional new approaches in “biasing” the rhizosphere. Fourth, the research community has just begun to reexamine their current strains and knowledge without an agronomic bias [e.g., Trichoderrna harziunurn, a “rhizosphere competent” fungal strain considered for biocontrol purposes that can degrade DDT, dieldrin, endosulfan, pentachloronitrobenzene, and pentachlorophenol in the laboratory (Katayama and Matsumura ( 1993)J.
1. T h e Rhizosphere a. Structure and Function The root-soil interface or “rhizosphere” is a zone of increased microbial growth and activity under the inHuence of the plant root. This zone is distin-
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guished from bulk soil by the root influence. Large microbial populations in the rhizosphere are sustained by exudation of carbohydrates and amino acids from the root and decortication of the root. These root cortical cells and root exudates are used as nutrients by microorganisms surrounding the root. Once established, microbial populations may be passively nourished by root exudation and decaying plant matter. The presence of microorganisms also may induce liberation of additional and/or specific organic molecules. Microorganisms in the root zone can also act as sinks for exudates, with the absence of bacteria leading to less production of this material (Campbell, 1985). First described for legumes by Hiltner, the rhizosphere has been the focus of agricultural research for many years, primarily because of its influence on crop productivity. Several excellent comprehensive reviews on the rhizosphere are available (Brock, 1966; Gray and Parkinson, 1968; Campbell, 1985; Curl and Truelove, 1986); therefore, only a brief description of important rhizosphere characteristics is presented below. Continual change at the root-soil interface, both physical and chemical, produces constant alterations in the soil structure and microbial environment. Differences between rhizosphere and bulk soil exist in carbon dioxide and oxygen concentrations, osmotic and redox potentials, pH, and moisture content (Foster et al., 1983). These parameters are further dependent on the properties and characteristics of specific plant species. Although modification of the soil environment in the rhizosphere by plant root secretions is an important process that influences microbial populations, the structure of the plant root provides microorganisms in the rhizosphere with an additional large colonizabie surface area. Fibrous root structures of grasses provide a larger surface area for colonization than taproot systems (Atlas and Bartha, 1993). The rhizosphere characteristics briefly described above translate into typical microbial populations which are an order of magnitude or more above microbial populations in nonvegetated soil. Using both light and electron microscopy, various researchers have described this increased abundance of microorganisms in the rhizosphere. This “rhizosphere effect” is often quantified as the ratio of microorganisms in rhizosphere soil to the number of microorganisms in nonrhizosphere soil, the RIS ratio (Katznelson, 1946). RIS ratios from 5 to 20 are common, but occasionally are as high as 100 and above (Katznelson, 1965; Gray and Parkinson, 1968). In addition to the increase in numbers of microorganisms associated with the plant root, the actual composition of rhizosphere microorganisms is dependent on plant species, plant age, and soil type (Campbell, 1985) as well as other factors including plant exposure to xenobiotics (Sandmann and Loos, 1984; Abdel-Nasser et al., 1979; Abueva and Bagaev, 1975; Gavrilova et al., 1983). Generally, the rhizosphere is colonized by a predominantly gram negative microbial community (Atlas and Bartha, 1993). The interaction between plants and microbial communities in the rhizosphere is a complex relationship. As indicated previously, plants sustain large microbial
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populations in the rhizosphere by rhizodeposition (Paul and Clark, 1989). Root cap cells, which protect the root from abrasion, may be lost to the soil at a rate of 10,OOO cells per plant per day (Campbell, 1985). In addition, root cells excrete mucigel, a gelatinous substance that is a lubricant for root penetration through the soil during growth. Soluble exudates include aliphatic, aromatic, and amino acids and sugars. In return for receiving exudate, microorganisms in the root zone can help solubilize insoluble nutrients and recycle organically bound nutritive elements. Ultimately, rhizosphere microorganisms are in a position that influences both loss of materials from roots and nutrient uptake by roots. The two most widely studied areas of mutualistic associations between microorganisms and higher plants are mycorrhizae and the legume-rhizobia symbiosis. Mycorrhizal infection of plant roots results in increases in nutrient uptake and is the primary mechanism by which many land plants absorb soil-derived substances. In addition, infected plants seem to grow better in nutrient-poor soils and there is growing evidence that mycorrhizal infections help plants to survive harsh environments including heavy metal-contaminated soil. In turn, the infecting fungus receives all of its energy requirements from root exudate (Allen, 1992). Mycorrhizae act as root extensions into the contaminated soil. Their length, small width, and ability to colonize soil aggregates greatly increase the root-soil contact. Such infections may increase the metabolic capacity at the root-soil interface. The fixation of atmospheric nitrogen is performed by a variety of autotrophic and heterotrophic microorganisms. The obligate association between leguminous plants and rhizobia begins with the excretion of exudates from the root to stimulate bacterial growth. While the stimulation is not limited to nodule-forming bacteria, legumes may stimulate the rhizosphere population more than nonleguminous plants (Nutman, 1958); however, a direct comparison of the effect of legumes and/or the process of nitrogen fixation on degradative capacity in the rhizosphere has not been done. b. Degradation in the Rhizosphere The excretion of root exudates by plants contributes to an increase in the density of microorganisms found near the plant root. In addition, the metabolic activity of microorganisms found living near plant roots is often greater than microorganisms in bulk soil. Such conditions provide for the possibility of accelerated microbial degradation in the rhizosphere and may be responsible for the increased degradation of organic compounds previously observed (Hsu and Bartha, 1979; McFarlane et al., 1981). In addition, the synergistic interaction between microorganisms may be required to completely degrade recalcitrant organics (Lappin et al., 1985). The specific use of plants to enhance microbial degradation of unwanted organic compounds in soils is a natural extension of these previous observations.
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The variety of compounds for which the enhanced degradation phenomenon has been observed suggests that plants might be managed at hazardous waste sites to facilitate the degradation of xenobiotics by rhizosphere microorganisms (Anderson et al., 1993; Shimp et a/., 1993). The use of plants may be particularly important when the organic contaminant is degraded “cometabolically” (not deriving energy or benefit from the degradation event), as the rhizosphere provides an environment conducive to cometabolic transformations (Hsu and Bartha, 1979). The next step in the use of rhizosphere microorganisms for enhancing microbial degradation of organic compounds in soils is to attempt to control the process. As root exudates are one of the most important components of plant! microorganism interactions, it seems logical to start with exudation as a potential controlling point. Root exudation (both composition and amount) can be dramatically influenced by plant species and age (Rovira, 1959) and plant nutrient status (Bowen, 1969). Environmental factors controlling root exudation are many (Gregory, I988), and include light intensity and temperature (Rovira, !959), soil moisture (Katznelson et al., 1955; Davies and Zhang, 1991), support medium (Boulter et af., 1966; Barber and Gunn, 1974), fertilization and exogenously applied chemicals (Wojtaszek et ul., 1993), and presence of mycorrhizae (Sundaresan et a l . , 1993). While the overall effect of root exudation is an increase in the numbers of microorganisms, the composition of the microbial community in the rhizosphere is also influenced by root exudate. The effect of such manipulations is further discussed (by contaminant class) in Section VIII. The choice of plant, manipulation of plant root exudates, and subsequent effect on the rhizosphere community are challenging and complex. It is hoped that as more research is completed in this area, we will acquire the knowledge and tools to: ( I ) further design, select, or engineer better plants; (2) more appropriately manage agronomic parameters to accelerate phytoremediation; and (3) make predictions of efficacy, efficiency, and relative achievable remediation goals. There is considerable room for innovation, basic knowledge development, and field work.
2. Endophytes To the chagrin of plant tissue culture experts, many plants have naturally occurring endophytes. These microbes within the plant tissue range in their effect on the plant from detrimental (and pathogenic), to commensual, or even beneficial. Certain of these microbes are even intentionally inoculated into seed on a commercial basis (e.g., endophyte-enhanced tall fescue). One such endophyte has been the target of a substantial amount of commercial activity. Its promise, although fading in agricultural relevance, may still be bright in a remediation sense as a method of delivering genes into the phytoremediation system.
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The basis of Crop Genetic International’s efforts in their h i d e technology has been the following: ( I ) molecular biology has isolated potentially interesting genes that may be beneficial to incorporate into plants; (2) plants currently have three genomes (chloroplast, mitochondrial, and chromosomal), none of which are as amenable to the tools of molecular biology as microbial genomes; (3) it is entirely possible to add a fourth microbial genome into the plant; and (4)monocots are a good place to begin as they are difficult to transform and regenerate by other methods. Based on these principles the nonpathogenic xylem endophyte Clavibacter xyli was isolated, altered to contain an insecticidal protein, and reintroduced back into the xylem streams in corn, rice, bermuda grass, etc. Although its presence sometimes resulted in a marginal decrease in crop yield that was agronomically not advantageous, the principle has been proven. The use of such a system in phytoremediation, where yields are not an issue, is still in its technical infancy. The use of xylem endophytes (either engineered or selected) is an intriguing possibility. An engineering technique employed today is the pumping of groundwater and/or slurried soil into an above-ground bioreactor. The possibility of a plant as a “self-contained, solar-driven bioreactor” may be technically achievable. The use of endophytic microorganisms for bioremediation appears to be good technology for augmenting the inherent metabolic limitations of plants. In addition, because of their internal association with plants, endophytes have a competitive advantage that free-living microbes do not. This close relationship should give both root and xylem endophytes a better chance for survival. In addition to xylem endophytes, root, leaf, and subcellular endophytes also exist. It is apparent that phytoremediation research may also require the talents of plant pathologists, virologists, and mycologists. Ectomycorrhizal fungi (Donnelly and Fletcher, 1995) appear to be the most likely candidates for this approach in the root zone.
3. Phyllosphere The leaf surface, or phyllosphere, is generally inhabited by microorganisms (bacteria, yeast, and fungi) adapted to withstand changes in sunlight, moisture, and temperature. The various bacterial and fungal populations which occupy the phyllosphere are largely dependent on leaf age, with highest numbers encountered close to the meristematic tissue. As the nutritional conditions on the leaf are dramatically different than those in soil, the composition of the microbiota on leaf surface is distinct and differs from the microbial composition of soil (Windels and Lindow, 1985). The metabolic capabilities of phyllospheric microorganisms may prove useful in removing organic contaminants from the air after deposition on the leaf surface, or after translocation from the root to shoot. Nonlethal fungal infections can also have dramatic effects in increasing transpira-
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tion rates (and hence pollutant flux) by by-passing stomata1 control over water flux. It also may be entirely possible to increase degradative capacity in the plant system by allowing a low-level, systemic/epiphytic, microbial infection to exist. As phytoremediation is not aggressively coupled to plant yield, it has been suggested that a simple plant-microbial relationship on the leaves (e.g., the nonaggressive powdery mildew strains associated with many woody dicots) may provide additional degradative/metabolic capacity in a phytoremediation system. True phyllospheric microorganisms may be more useful in aquatic environments where conditions are less extreme. In these environments they exist in biofilms on the surface of aquatic plants. Federle and Ventullo (1990) reported increased microbial degradation of surfactants by microorganisms associated with leaves from a laundromat wastewater pond. Mineralization of the surfactants (linear alkylbenzene sulfonate, linear alcohol ethoxylate, stearykrimethyl ammonium chloride, and distearyldimethyl ammonium chloride) occurred without a lag phase and degradation half-lives for all of the surfactants tested were less than 14 days.
C. PLANT-PRODUCED PHYSIOCHEMICAL EFFECTS Soil scientists, and those that study root physiology, speak about a set of relatively intangible parameters such as “tilth,” “nutritive status,” “health of the soil,” and “sustainable soil quality.” These terms have imprecise technical definitions, and there is a set of political, regulatory, and societal questions associated with their various definitions. Beyond the rhizosphere effect, the effect of root exudates and enzymes, and the uptake and degradation of toxins by plants, the growth of plants seems to affect these “soil health” parameters. In our experience, the growth of certain hardy plants in a contaminated soil can ameliorate the soil for growth by other less hardy plants. Experiments with petroleum hydrocarbon contaminated soils in our labs (S.D.C., F.C.H.) have shown that a clay soil with up to 3% TPH and extreme phytotoxicity to all crop plants tested could be planted in a tolerant grass species (Vefiveria zizanioides) which thrived. (Vetiver grass is a perrenial grass grown widely throughout the world for erosion control; Vietmeyer and Dafforn, 1993). After a 1-year period of growth, other species could also be grown in that soil; although there was no detectable change in quantity or spectrum of contaminants over the test, bioassay results on plant growth showed considerable differences. We are not converts to the Gaia movement; however, it is obvious from field experience that if and when plants can be grown in soils where they are not currently present, the soil can, by careful husbandry, become less of a biological
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risk. Plants alter respiration rates, and add carbon and other nutrients to the soil. They can increase total microbial activity, accelerate nutrient cycling, increase organic matter, form and stabilize secondary and tertiary soil structure (e.g., aggregates and peds), increase water infiltration rates, and move water and nutrients in three dimensions. Agronomic practices such as tilling decrease field heterogeneity and increase contact and transport between microbial degraders, contaminants, and cometabolic substrates. All of these processes are postulated to decrease the potential negative impacts that a contaminant would have on the environment. As one engineer suggests, “ln the final analysis growing plants couldn’t hurt. . . . could it?’
VI. MODELING PHYTOREMEDIATION Most models of phytoremediation are attempts to utilize models designed for other purposes. Plant and root growth, water uptake by plants, and solute transport are all commonly modeled by agronomists (Hanks and Richie, 1991). In addition, models also exist for general uptake of xenobiotics into plants (see Section 1V.A; Hathway, 1989), as well as the larger environmental fate of these materials: PRZM and PRZM-2 (Pesticide Root Zone Model-Carsel et a l . , 1984). Other models are larger still in scale but might be adaptable to phytoremediation. These include the “Mackay” model (Mackay, 199 I ; MacKay et al., 1992) and “Simple Box” (Van de Meet 1993). Other models that may also be useful as starting points in the development of phytoremediation include SoilFug (a soil fugacity model), LEACHM (solute transport model in agricultural soils), PELMO (pesticide leaching model)m HAZCHEM (evaluation of chemical fate for hazard assessment purposes), and the soon to be published CHEMCAN (Efate model). No model exists, however, that has the relevant input and output parameters that the remediation community seeks. Input parameters should include concentration, bioavailability, rhizosphere degradation, plant uptake and degradation, and harvesting. Output should include all the technical parameters and sensitivity analysis suggested by other models, and at the same time should answer some of the more fundamental questions encountered at the site level. These questions include: How long will a remediation take? What are the variables that it depends on most? What endpoint can we reasonably expect to obtain? We believe that additional efforts in modeling the process of phytoremediation are appropriate. Since much of the remediation community is “cost-engineering’’ conscious, a cost-predictive capacity in at least some rudimentary model would be helpful. Cost models should include capital, operating, and “net present”
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costs. Site managers could probably supply useful estimates on such unfathomable parameters as “lawyer hours per ton remediated” or alternatively “legal council required per useful event.” Many of the existing models rely on physical and chemical properties of the chemicals and soils including KO,, K,, water solubility, volatility, soil texture, pH, and soil organic carbon content. Most often modeling uptake of xenobiotics by plants involves a series of mathematical representations including equilibrium and the kinetics of partitioning of the compounds within the soil and plant, diffusion, mass flux, and plant metabolism of the xenobiotic (Trapp, 1995). In phytoremediation it should further include such factors as humification, degradation by microbes, and volatile loss from leaf surfaces. Because experimental data for many of the compounds are lacking, any model must rely on many estimates. For example, it is assumed that log KO, may be used to predict water solubility, sorption to soil, mobility in the xylem, and mobility in the phloem. Most of the common chemical and physical parameters are available for a wide range of pollutants (K,,, Henry’s law constant, molecular size, etc.); however, the biological parameters in interactions require either actual measurements or estimates obtained from other literature. More work and data generation are needed in the area of modeling; however, models are useful not only in their predictive capacity, but also because the process of model-building requires building a set of relationships and their relative degree of importance, which can point out fundamental unknowns in phytoremediation as well as their importance or triviality. In addition, models of all sorts are extremely useful in discussions with regulatory agencies and the engineering community.
VII. PRACTICAL CONSIDERATIONS A. SITECONDITIONS AND LIMITATIONS Most sites in need of remediation are not ideal “farming sites.” Significant costs can be incurred in initial site preparation. Abandoned manufacturing sites, terminals, and parking lots may present no added burden to engineering techniques, but plowing, rototilling, ripping andlor seeding equipment generally works poorly in these areas. In some areas where debris removal or a “size reduction” program is impractical, “low till” methods become increasingly important. Hand-seeding has been necessary in some cases. Urban waste sites may provide their own challenges. Phytoremediation relies on living plants and is therefore sensitive to other environmental stress. In our experience metals (Cr, Pb, Hg, Zn, Ni, Cu, Cd) and
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salts (Na) are the most commonly encountered on sites contaminated with organics. Adjustments in plant selection, water regime, soil pH, and planting system generally can be made to alleviate some of these problems; however, certain contaminant profiles preclude phytoremediation. We also encountered field sites that were “on the average” only barely contaminated. There was, however, a 10,000-fold variation between samples. Most soil scientists are familar with the heterogeneous nature of soils. Soils at hazardous sites, however, are often considerably more heterogeneous than those found at agronomic sites. Sampling, field layout, and statistical design are difficult when variability is high. Engineers often work on the sampling unit of a bucket loader. Such scale considerations must be considered in phytoremediation as well. Some petroleumcontaminated sites have visual tar balls. Much forethought should be given to the experimental design at such sites. In our experience it is beneficial to specifically address: (1) sampling techniques and sample sizes; (2) homogenization, extraction, and analytical protocols; and (3) personnel sampling predilections (a preoccupation or aversion to tar balls being included in the sample). We have had more than one uninterpretable experiment due to these field and sampling considerations. The regulatory status of the site sometimes provides additional unique constraints. Our experience at one site may be enlightening. Although the site is considered “hazardous,” the pollutant is not legally classified as “hazardous waste”. That definition only applies to attempts to dispose of it. At this site, a “hazardous waste” was created, however, in the process of our digging holes to plant tree seedlings. Our shovel-fulls of “hazardous material” became “hazardous waste” when we lifted them above the plane of the earth. As returning the soil to around the plant roots was a “land disposing of a known hazardous waste” (which requires a permit “to operate a land disposal unit” that is no longer granted in that state) all planting was halted. Planting resumed, however, on the advice of legal council, when holes for planting were dug with a gasolinepowered, rotary agar and soil was pushed back into the holes around the plants “in a horizontal motion so as not to break surface contact.” Interestingly enough, legal opinions may also be obtained on mold board plows, rototillers, and handheld weeding devices. Most researchers attempting to “farm” hazardous waste sites have similar stories of legal (or site guidance) procedures preventing even the simplest of farming tasks. Those working on contaminated sites may be well advised to expect such complexities. The decontamination of a tractor can be an exceedingly complex task when supervised by a full complement of industrial and regulatory lawyers (one author’s personal experience). i n addition the term “plant staff” generally refers to the manufacturing facility and does not indicate any experience or capability with vegetation management. Field work in this area can lead to interesting cross-cultural experiences. One soil microbiologist had to be
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“reminded” that it was not necessary to “flame” his shovel prior to taking soil samples. This was particularly noteworthy in the vicinity of large gasoline spills. If we had to choose an ideal phytoremediation site it would be a rural agricultural field, moderately contaminated with recently spilled crude oil, with no possibility of contaminating the surface water with runoff. Crude oil qualifies under the petroleum exclusion of the Resource Conservation and Recovery Act (RCRA) and, thus, eliminates many regulatory constraints. Rural agricultural sites tend to be less regulated due to reduced exposure and easier control. A selfcontained field with little chance of runoff further reduces risk of exposure to the contaminant; unfortunately, few sites are this ideal. Because contamination rarely occurs in ideal settings, attempts at remediation are filled with challenges. Regulations imposed on hazardous waste and contamination may require installation of berms to contain runoff, the collection and recycling of leachates, groundwater monitoring, and analysis of above-ground plant parts to ensure that hazardous chemicals are not accumulating. One site required catching muskrats for fat tissue sampling. These constraints may require special engineering approaches in order to implement all phases of the cropping systems. Sampling, tilling, and seeding may require special modifications. Special no-till seeding equipment (or hand seeding) and regulatory variance may be necessary. If toxic chemicals are accumulated in the upper portions of the plants, the system may be further complicated: grazing animals may have to be excluded from the site, and disposal of the plant material may become a regulated event in the event that it cannot be disposed of on site. Although this has not been our experience to date, composting, landfilling, or incineration of the plant material may be necessary in certain instances and may significantly add to the costs.
8. AGRONOMICS Plant species that are chosen must be well adapted to the climate of the region irrespective of varietal traits specific for phytoremediation. Length of growing season, average temperature, and rainfall are important considerations. Species that perform well near the Gulf of Mexico are not likely to perform well in the Rocky Mountain region or in desert areas. Plants used on the Gulf coast will need to be heat tolerant and should take advantage of the very long growing season. In higher elevations, the growing season is shorter, and freeze tolerance may be more important. In regions that can have extended dry periods without the advantage of irrigation, drought tolerance will be critical. Local experts should be consulted about the proper plant varieties to use prior to initiating field trials for phytoremediation. County extension agents, USDA plant materials centers, and local seed dealers are often excellent resources for this information. Choos-
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ing an inappropriate species or variety could result in a failure or a negative bias in the results. There are other considerations involving the physiology of the plant. Plants utilizing the C4 photosynthetic system can tolerate heat and drought better than C3 plants and are more efficient in the use of CO,. Many of the C4 grasses also are obligate mycotrophs; therefore, a comparison of C3 versus C4 plants would have merit. Proper germination of the seeds can be ensured only if the soil is moist (but not saturated), the temperature is ideal, and the soil is not toxic to the seeds or seedlings. Phytotoxicity assays are almost always carried out prior to planting; however, they may not be necessary when the site is already partially to fully vegetated. Irrigation or drainage may be necessary to control the soil moisture, and soil physical properties may dictate management of these operations. Soil temperature is important for germination of seeds and is most practically controlled by date of planting. It may not be possible to control the date of the initial planting. Regulatory or management requirements may demand that the field be planted at a time that is not ideal for the plants best suited for phytoremediation in the area. In our experience, plant personnel and regulators enjoy seeing a Ph.D. dressed up in Tyvek and respirators on the hottest day in July planting the “spring crop” on a contaminated site. Alternatively, “fall” planting can occur at similarly nonideal times. If planting at less than ideal times is required, a cover crop that can be planted to stabilize the soil surface should be considered. Nitrogen-fixing plants, such as the legumes, have been used in some trials because nitrogen is a critical component in the mineralization of organics in soil. Legumes provide their own nitrogen and microbes need not compete with the plant roots for N during mineralization. Soils contaminated with high levels of many organics are often inherently N limited due to rapid microbial degradation. The addition of larger than customary amounts of nitrogen or the planting of legumes may be beneficial. Phytoremediation is still in its nascent stage. No systematic study of plantspecific traits has been carried out to elucidate generally important parameters. To test the hypothesis that a property of a plant enhances phytoremediation, one would test two varieties that are identical in every way except for that property. For example, N, fixation is viewed as a property with the potential to enhance degradation. Some species of legumes have pairs of varieties that are identical except that one variety will lack the ability to associate with the appropriate bacteria so as to fix nitrogen. The importance of this trait can then be tested in a simple greenhouse trial using these paired varieties. The systematic selection of traits to enhance phytoremediation is still in its initial phase. More work in this area is required. Once these traits have been better identified, however, their incorporation into varieties of plants that are
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suited for the region of interest is relatively straightforward. Several techniques are available for this process. 1 . Plant breeding: Crosses and/or selections are made to build up a seed source that has the desired property or combination of properties. 2. Tissue culture-cellular biology: These techniques operate at a cellular level rather than on the whole-plant basis of classical plant breeding. 3. Biotechnology: If genes that are responsible for a certain trait can be identified, then the incorporation of these genes into more desirable varieties is possible with the multiple transformation, selection, and regeneration techniques currently in use. Biotechnological approaches are not universal. They are gene and plant specific,sometimes at the cultivar level (e.g., corn). The transformation and regeneration of some species is still beyond our technical capacities; however, similar manipulations on other species have become routine.
The objectives and approaches to establishing a vegetative cover for phytoremediation are considerably different than those for crop production, reestablishment of native species, or simple reclamation of disturbed sites. Typical plant uses focus on the parts of the plant above ground in order to optimize grain yields, maintain long-term stands, or establish maximum cover. Phytoremediation of organics is oriented primarily to the below-ground processes, high root biomass, enhanced microbial activity, and dissipation of organic contaminants. Plant establishment for phytoremediation follows many of the standard procedures for land reclamation or crop establishment, but with some modifications. First, varieties must be adapted to the climate of the region. However, the final selection of species and varieties will be based on maximum absorptioddegradation of the organic contaminants, which is discussed further below. Second, seed beds must be properly prepared which may pose logistic difficulties in urban, manufacturing, refinery or terminal settings. As with normal practices, the soil may be tilled, weeds controlled, and seeds or transplants planted at their proper depths. For phytoremediation, plant spacing (seeding rate) will be adjusted to maximize root density, not to attain maximum yield. The soil may have some unfavorable chemical and physical properties as a result of the contamination that needs to be overcome prior to planting. For example, spills of crude oil may also be associated with high concentrations of soluble salts, particularly sodium. Special management of the salt may be required before seeding. Third, nutrients should be monitored and applied as needed. Because the objective of bioremediation is degradation, nutrient levels will be adjusted for optimum microbial activity, not maximum return on fertilizer investment. This
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generally will require adding much more nitrogen and phosphorus than needed for the plants alone as the process of active microbial degradation also requires nutrients. Fourth, established plants need to be managed to ensure maximum growth. Moisture levels must be optimal, weeds may need to be controlled, and certain plants need routine mowing.
C. LIMITATIONS TO PHYTOREMEDIATION Phytoremediation is not a universally applicable or perfect system. It offers many positive aspects, but there are drawbacks as well. Not all contaminants or contaminated soils will be suitable for phytoremediation. Bioavailability issues, contaminant leachability concerns, heterogeneity of the site, and complexity of the contaminant profile may severely restrict many, if not all, forms of phytoremediation. Other considerations that have also been mentioned are legal constraints, the logistical need for advanced planning (which may not coincide with weather changes), and access to the site. There are significant depth limitations to phytoremediation. Although some roots have been reported at a depth of 60 m, most roots are relatively shallow. Researchers working with relatively mobile contaminants and long-lived perrenials (e.g., trees and some grasses) believe that contaminants as deep as 10 m could be drawn up to the plants by a combination of the root presence and the hydraulic gradient induced by transpiration. With relatively immobile contaminants, most researchers consider remediation possible only within the top 0.5 to 1 .O m. Although the maximum depth to which roots can be found will often be considerably deeper, the root density decreases dramatically with increasing depth (Foxx et uf., 1984). Contaminants that cannot be drawn to the plant roots over long distances are increasingly unlikely to be affected by phytoremediation as depth increases beyond 1-2 m. All plant roots require oxygen for respiration. Under conditions where soil texture, water content, or microbial respiration rates are high, roots may not penetrate. There are some plants, most notably plants accustomed to waterlogged situations, which have adapted to such environments. These plants have specifically adapted structures in their stems and roots to transport oxygen from the air to the root zone. Most terrestrial plants, however, die under anaerobic stress. In areas of high organic contamination rapid microbial degradation decreases the partial pressure of oxygen below the point where many plants can survive. These anaerobic zones are not appropriate targets for phytoremediation until the microbial degradation (and its effect on oxygen consumption) has subsided, Areas with insufficient oxygen concentrations due to restrictive oxygen flux (water logged, too tight a soil texture or profile) can be targeted after
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alleviating the 0, influx rate limitation by engineering or agronomic manipulation. Compounds present at phytotoxic concentrations cannot be treated unless a resistant species is selected or the soil is pretreated to reduce phytotoxicity. Fresh crude oil and high levels of trinitrotoluene (TNT) are known to inhibit plant growth. The phytotoxic agents in the crude oil appear to be volatile compounds that disrupt cell membranes; simple aging (or “weathering”) of the contaminated soil reduces phytotoxic effects due to dissipation of the volatiles. Phytotoxicity of TNT is apparent at soil concentrations as low as 30 mg/kg. Initial use of active or intrinsic bioremediation to reduce concentrations will encourage the establishment of vegetation and phytoremediation. Traditional ex situ engineering methods are relatively rapid. Traditional bioremediation (landfarming) of oily soils is slower and may require a number of years. Phytoremediation may require even greater periods of time. In an economic sense, the trade-off may be between time involved in remediation and cost of the remediation. If time is a limitation (due to land use considerations or exposure potential), it may always be faster to remediate the site by engineering techniques. The one exception is soils at such a large scale that engineering scale-up is impossible (e.g., incineration of 10,000 acres may require decades, as much time as the longest phytoremediation projections). Although the use of managed ecosystems to restore damaged soils is appealing, there may be certain detrimental ecological effects that have to be managed as the remediation occurs. Assimilation of the pollutant by plants is viewed as desirable in phytoextraction; however, the potential for food chain contamination must be carefully considered. Volatilization of contaminants into the atmosphere may also present an unacceptable problem if compounds are absorbed, translocated and transpired. None of these ecological concerns seem to be insurmountable by careful system design, engineering controls, or better selection of plant species. During the early phases of plant establishment the soil will be bare and subject to erosion. Contaminants adsorbed to the eroding soil or dissolved in the runoff water are potential ecological threats. Erosion control mechanisms, similar to those commonly used at contaminated sites undergoing other forms of remediation, may need to be employed. In some cases increased leaching of soluble compounds might occur as the soil texture is loosened by plant growth. The presence of plants, particularly perennial species, might enhance infiltration by creating channels from the surface to the extent of the roots. As most phytoremediation projects do not target sites with large amounts of water-soluble components such a case is thought to be rare, but should be examined. Water management schemes may help mitigate any such questions or problems that might arise.
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VIII. CURRENT PHYTOREMEDIATION RESEARCH AND DEVELOPMENT A. PETROLEUM CONTAMINATION April and Sims (1990) studied the effects of using deep-rooted prairie grasses to remediate soil contaminated with polycyclic aromatic hydrocarbons (PAHs). They suggested that the roots of these perennial grasses may be more effective at stimulating the rhizosphere microflora due to their fibrous nature. These fibrous roots offer more root surface for colonization and soil/root contact and can explore soils up to depth of 3 m. Sod-forming grasses are also excellent soil stabilizers and could reduce erosion as well as remediate waste sites. April and Sims (1990) investigated the use of big bluestem (Andropogon gerardii), indian grass (Sorghasrrum nutans), switchgrass (Panicurn wirgatum), Canada wild-rye (Elymus canadensis), little bluestem (Schizachyrium scoparius), side oats grams (Bouteloua curtipendula), western wheatgrass (Agropyron smithii), and blue grams (Boureloua gracilis). Manure amended soil was contaminated with benz(a)anthracene, chrysene, benzo(a)pyrene, and dibenz(a,h)anthracene. It was then seeded with a mixture of these prairie grasses. After 219 days of growth, PAH dissipation ranking among the four PAHs studied correlated with the water solubility of the compound, with the most soluble compound showing the highest degradation. Schwab and Banks (1994) investigated the degradation of PAHs in the rhizosphere of a variety of plants grown in landfarmed, petroleumcontaminated soil. Degradation in the contaminated soils was assessed for different plant types and PAH compounds. The plants used in this study included alfalfa (Medicago saliva), fescue (Festuca arundinacea), big bluestem ( A . gerurdii), and Sudan grass (Sorghum vulgare sudanense). The PAH compounds evaluated were anthracene and pyrene. Their results also indicated that PAH compounds dissipated at a faster rate in rhizosphere soil than in unvegetated soil. As of 1995, a half-dozen field tests and as many lab tests are underway to further quantify this rate of degradation relative to traditional landfarming techniques and the static unmanaged condition. Much more work remains to be done in plant selection and management techniques to maximize this effect.
B. CHLORINATED HYDROCARBONS 1. Fthizosphere Degradation
As a class, chlorinated hydrocarbons represent some of the most persistent chemicals known. Although there are many halogenated organics produced natu-
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rally, they are not common constituents of most soils and may be resistant to biodegradation primarily because of their insolubility, molecular size, toxicity, and/or inherent chemical bond energies and configuration. Although use of these compounds as pesticides or industrial chemicals has dramatically decreased or ceased in many countries, their persistence has contributed to their continued detection in the environment and occurrence at hazardous waste sites. Despite this persistence, the degradation of some chlorinated hydrocarbons has been shown to be enhanced in the rhizosphere or by microorganisms from rhizosphere soil compared to nonvegetated soils. Pentuchlorophenof (PCP) is a widespread chlorinated hydrocarbon contaminant, especially in soil around wood treatment plants. At high concentrations, PCP is often quite persistent due to its toxicity to a variety of soil organisms. One approach to remediation of PCP-contaminated soil has been inoculation with microorganisms capable of metabolizing PCP (Crawford and Mohn, 1985);however, survivability of inoculated organisms is often limiting. Ferro and coworkers ( 1994) explored the possibility of using rhizosphere microorganisms to accelerate PCP degradation. Crested wheatgrass (Agropyron desertorurn) was tested for the ability to enhance mineralization of 14C-PCP in a flow-through soil-plant system in the laboratory. Mineralization of 1%-PCP in the vegetated system was 22% of the applied 14C after 155 days, whereas only 6% of the 14CPCP was mineralized in the nonvegetated system. A significant portion of the I4C-PCP and/or metabolites (36%)was taken up by the plants. Leaching of 14CPCP and/or metabolites was greater in the nonvegetated system. Overall, the vegetated soil had lower levels of PCP-derived material than the nonvegetated soils at the conclusion of the experiment. Thus, vegetation was beneficial in increasing mineralization of PCP as well as reducing leaching of PCP and its metabolites. Trichloroethylene ( I , I ,2-trichloroethylene) (TCE) is one of the most common substances found at hazardous wastes sites and in contaminated groundwater. Walton and Anderson (1990) previously observed accelerated degradation of TCE in slurries of rhizosphere soil collected from four plant species at a former solvent disposal site and enhanced mineralization of I4C-TCE in rhizosphere soil samples. However, the authors speculated that the absence of a living plant may have lead to conservative estimates of degradation for TCE, and in subsequent experiments utilizing soil-plant systems composed of soil and vegetation from the same contaminated field site showed significant mineralization of 14C-TCE. In this study, specially designed Erlenmeyer flasks were used to monitor the fate of “T-TCE in the soil-plant systems. In addition to the vegetated samples, nonvegetated and sterile soil samples were included as separate treatments for each experiment. The flasks were sealed by coating a nonreactive silicone rubber sealant between the plant stem and the flask. At 24-hr intervals, headspace within the flasks was flushed through a series of traps for removing IT-volatile com-
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pounds and W02.Analysis of the I4CO2traps from the experiments showed that I4CO2production in the vegetated soils was elevated compared with both the nonvegetated soil and the sterile control soil. In the experiments with soils containing Lespedeza cuneatu and Pinus tuedu, I4CO2production at the conclusion of the experiment was significantly greater than I4CO, production in nonvegetated and sterile (autoclaved) control soils (Anderson and Walton, 1992, 1995). Most of the 14C02produced from 14C-TCE in the vegetated soils evolved during the initial 3 days of the experiment. This is in agreement with earlier observations on the initially rapid disappearance of TCE from the headspace above aqueous slurries of rhizosphere soil (Walton and Anderson, 1990). In addition, comparisons of the percentage of I4C-TCE mineralized in the whole plant systems with previous data on “T-TCE mineralization in L . cuneata rhizosphere soil appeared to confirm the hypothesis that the mineralization rates based on rhizosphere soil samples gave conservative estimates of mineralization that would occur in soil containing a living plant. Polychlorinated biphenyls (PCBs)provide another research story of particular interest in the relationships between rhizosphere microorganisms and plants. While the use of microorganisms for remediation of PCB-contaminated environments has been the focus of research for the last 10 years, recently the use of vegetation for enhancing microbial degradation has also been explored (Donnelly et a / ., 1994; Donnelly and Fletcher, 1995; Brazil et a / ., 1995). The use of ectomycorrhizal fungi for bioremediation is a potential technology for overcoming some of the survival limitations of soil inoculation. The symbiotic relationship of plant-ectomycorrhizal systems may give the fungus a better chance to compete against indigenous soil microflora and subsequently increase PCB metabolism in contaminated soils. Donnelly and Fletcher (1995) explored the metabolism of a PCB stock solution containing 10 or 11 different PCB congeners (2-6 C1) by ectomycorrhizal fungi in culture flasks over 5 days. Of the 21 fungal species tested, 14 were capable of some PCB metabolism. The number of congeners metabolized ranged from 0 to 7 and varied among the fungal species. As expected, the lower-chlorinated congeners were more easily metabolized. Radiigera atrogleba and Cautieria crispa metabolized the most PCB congeners. An interesting note familiar to many in the bioremediation area is that there was no correlation between taxonomically related species and metabolism of structurally similar congeners. Donnelly and co-workers (1994) have also been exploring the use of PCBdegrading bacteria in combination with plants. Previous research from the group at General Electric (Bedard et a / . , 1986) identified several aerobic bacteria capable of PCB metabolism. However, success with these organisms in the field has been limited by the requirement of biphenyl as a cosubstrate. The need to find other compounds which stimulate the growth of PCB-degrading bacteria led Donnelly and co-workers to hypothesize that naturally occurring compounds
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produced by plants may be useful. The growth of three bacterial strains shown to degrade PCB was tested using several known plant compounds including flavonoids and coumarins. Growth on biphenyl served as a control. Several of the plant compounds supported growrh of the bacterial strains as well as or better than biphenyl. In addition, bacterial strains grown on the plant compounds retained the ability to metabolize certain PCB congeners. Results of these studies suggest that certain plant species or stages of plant growth might be valuable for enhancing PCB degradation in soil. Plants can select for different rhizosphere microbial communities. It remains unproven, however, whether this selection affects degradative rates in the field. Whether this selection potential translates into differences in the rates of microbial degradation of organic compounds remains to be proven. It is on this premise, however, that recent research by Fletcher et al. (1995) has screened and selected plants for specific exudate patterns and has examined the rate and timing of the release of these compounds into the rhizosphere environment.
2. Remediation in Plantu As mentioned above, research on pesticides has shown that plants have four known types of reactions which result in the breakage of the C-Cl bound. These reactions include: (1) monoxygenases, (2) glutathione (or homoglutathoine) S-transferases, (3) (anti)auxin cell receptor binding (that converts a C-Cl to a CS protein), and (4) a nonenzymatically catalyzed replacement of a C1 to an OH (in certain aromatic configurations, e.g., triazines). Other mechanisms of C-Cl bond breakage undoubtedly exist; however, they have not been actively researched. The capacity of plants to metabolize PCBs has been studied by Groeger and Fletcher (1988). The extensive screening on whole plant showed a number of differences in metabolic capacity between plants. To ascertain whether this was plant metabolism or a microbial-plant community, Groeger and Fletcher went further and produced and screened plant tissue culture (free from microbial associations) for PCB metabolism. Cell cultures of rose (Rosa sp., cv. Paul’s Scarlet) were found to have among the best ability to metabolize PCBs, but the completeness of metabolism was dependent upon specific PCB congeners. Recent work on trichloroethylene in hybrid poplar trees has shown that the plant is relatively insensitive to the presence of the compound and shows no ill growth effects at 100 mg/liter TCE (many times higher than the level deemed safe in drinking water). The TCE is taken into the plant and has two detectable fates: it is either metabolized or bound. The metabolic process produces 2,2,2trichloromethanol and di- and trichloroacetic acid compounds, which would suggest the presence of an active TCE metabolizing P450 enzyme (Strand et al., 1995). In addition, a large fraction of the TCE taken up by the plant roots
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becomes bound and unavailable to chemical solvent extraction. More work on this fraction is also needed from a mechanistic as well as an ecotoxicological perspective. Based on previous results, this bound fraction may also be unavailable to animals, microbes, and other environmental receptors. Binding pollutants into the woody tissue of long-lived trees may prove to be an acceptable phytostabilization strategy. A field test of these poplar trees has recently been installed by these same researchers.
C. PESTICIDES The widespread use of pesticides during the last 40 years has facilitated the growth of retail agrochemical dealerships. Unfortunately, many of these dealerships have experienced soil and water contamination problems from normal operating procedures and accidents. At pesticide-contaminated sites, a potential limitation to using vegetation exists because of the presence of mixtures of herbicide contaminants. Nonetheless, herbicide-resistant plants exist at these sites, and rhizosphere soils from these plants have previously shown the ability to degrade mixtures of herbicides (Anderson et u l . , 1994). In addition, previous studies on the herbicide-degradative capability of rhizosphere soils of other plant species (Sandmann and Loos, 1984; Lappin et a[., 1985) help support the use of vegetation in remediating pesticide-contaminated sites. The use of plants in the remediation process for these materials is also a logical extension of ongoing research in the landfarming (surface tilling and fertilization) of these same materials (Felsot and Shelton, 1993). 1. Rhizosphere Degradation
Anderson and co-workers have conducted studies utilizing soils and plants from pesticide-contaminated sites. Preliminary studies indicated increased degradation of atrazine, trifluralin, and metolachlor in rhizosphere soil from Kochia scoparia. a herbicide-resistant plant, compared with nonvegetated soil (Anderson et a l . , 1994). Subsequent experiments indicated that mineralization of 1%atrazine in soil treated with a mixture of atrazine and metolachlor at concentrations typical of point-source contamination (50 x g/g each) was significantly greater in rhizosphere soil from Kochia scoparia than in nonvegetated and control soils (Perkovich et a l . , 1995). Soils were collected from an agrochemical dealership contaminated with several herbicides, including atrazine, metolachlor, trifluralin, and pendimethalin at concentrations well exceeding the field application rates. Mineralization rates of ring-labeled atrazine in both rhizosphere and nonvegetated soils were quite high (>47% of the initial 14C applied after 36 days) compared to literature values. Based on the relatively rapid miner-
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alization half-lives of atrazine in both soils, it does not appear that the presence of metolachlor at 50 pg/g had a negative influence on the degradation. This research supports the use of rhizosphere microorganisms associated with herbicide-resistant plants to enhance microbial degradation of atrazine in soil at contaminated sites. Naturally occurring plants, such as K . scoparia, appear to have the capacity to be used as in sifu agents of bioremediation by facilitating the proliferation of microorganisms in surface soil with the ability to mineralize high concentrations of atrazine. Anderson and Coats (1995) have also screened other rhizosphere soils from waste areas for their ability to degrade atrazine and metolachlor. Several soil samples exhibited the ability to mineralize high concentrations of I4C-atrazine. These included rhizosphere soils from lambsquarters (Chenopodium berlundieri), foxtail barley (Hordeumjubatum), witchgrass (Panicum capillare), catnip (Nepeta cataria), and musk thistle (Carduus nutans). Of the 14 species (eight families) tested, the greatest mineralization of 14C-atrazine was observed in rhizosphere soils from musk thistle (33.1-1.7%) and catnip (24.1-1.2%). However, none of the 14 rhizosphere samples tested exhibited a positive response for 14C-metolachlormineralization.
2. Remediation in Plan@ Actually, modern pesticides, and in particular herbicides, are an ideal target for phytoremediation. The materials are designed for application on soil and plant systems. Herbicides which are designed to be applied to the bare soil at planting, or prior to planting, are specifically designed to move through the soil to the plant root, be taken up by the plant roots, and be metabolized by the tolerant species. In one sense, phytodecontarnination occurs in many farmers’ fields, as an overapplication of a pesticide may require more than one cropping cycle to reduce its bioactivity down to a level where sensitive species can be grown. The phenomena of “carry-over” is often considered good by farmers who grow successive monocultures of the same crop (as it reduces their need for reapplication), but bad by farmers who try to replant with sensitive rotational crops. Many farmers will relate stories of altering crop management schemes to accommodate a spill or miss-application (due to incorrect dilution procedures or malfunctioning equipment). One farmer, who had planted corn into a triazine spill a decade earlier, thought the Ph.D.’s “concept of phytoremediation of pesticides” was “about half a bucket of common sense” and then went on to inquire what we did for a living. The mechanism of triazine uptake, degradation, and characterization is exhaustively studied in the literature, and also apparently practiced with some skill in the field. It is not, however, a soil decontamination method approved by most regulatory authorities.
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D. BIOTECHNOLOGICAL IMPROVEMENTS IN PHYTOREMEDIATION Over the last decade, the transformation and regeneration of microbes and plants has advanced greatly. Genes are now commonly moved between microbial genera, and genes cloned from a wide variety of viral, microbial, plant, and animal sources are now routinely expressed in some plants. Genetic engineering for enhanced phytoremediation is in its infancy, yet progress is being made. Progress has been made in finding and cloning potentially useful genes, transforming and regenerating appropriate species, altering plant characteristic morphology, changing plant metabolism, and adding degradative capacity to the rhizosphere.
1. Degradative Genes A wide variety of microbial genes that can metabolize xenobiotics have now been cloned. A small portion of these have been expressed in plant tissue. The primary purpose behind most of this work to date has been to increase the tolerance of a crop plant to a particular herbicide. Such altered plants may increase the sales of the associated herbicides. Examples of this include: ( 1 ) Alzodef tolerance conferred on tobacco by the introduction of the cah gene from the fungus Myrothecium verrucaricr (Maier-Greiner et ul., 199l ) , ( 2 ) Glofosinate tolerance in tobacco and potato with the bar gene from Strepromyces hyet al., 1989), and ( 3 ) a glyphosate metabolism system that g r o s c o p i c ~(DeGreef ~ may be used in conjunction with the altered tolerance genes (again a microbial source) in “Round-up ready” germplasm developed by Monsanto. In addition to these, P450s have been cloned from microbial, plant, and mammalian tissues. Some of these, including mammalian genes, also have been expressed in plants (Saito et al., 1991). Many molecular labs are now in the process of isolating the genes responsible for a specific metabolic activity in microbial, plant, or animal systems. Insect populations resistant to specific insecticides have even provided a source of novel glutathione S-transferase activity (Thompson er a l . , 1994). These genes are being cloned and eventually expressed in plant tissues. In some cases, the molecular strategy has needed some adjustment (e.g., whole-plant, constitutive expression of a gene coding for the metabolism of a PCB is unlikely to succeed as PCBs are a lipophilic contaminant that, if available, are tightly bound into the outer layers of the plant root). Although the molecular biology may be of intrinsic interest, justifying the project on the basis of phytoremediation potential is difficult. The best use of molecular biological tools is as part of an integrated phytoremediation team. To a large degree, the limiting factors in phytoremediation are unknown and hence difficult to target with molecular biological skills. Relatively few recombinant plants have been made specifically for the purpose of phytoremediation, however the microbial genes chlorocatechol 1,2-dioxyge-
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nase, catechol 2,3-dioxygenase (Gordon et al., 1990), and chlorophenol hydroxylase (Stomp el al., 1994) have all been engineered into plants for that purpose.
2. Other Targets Much is known about the plant metabolic pathways that produce bioactive compounds. Of particular interest is the regulation of pathways that determine the plant’s production of microbial signal molecules, antibiotics, pigments, and xenobiotic pollutant analogs. The biosynthesis of many of these molecules is increasingly well understood (Kubasek ef al., 1992), and in some cases genes (e.g., PAL pathway genes) have been isolated and cloned into vectors appropriate for phytoremediation work. Work at this level would be more likely to produce research tools than to actually produce field plants at the moment; however, the prospects are intriguing. Producing paired plants differing in only one trait (e.g., 100-fold difference in tannin or phenolic exudation) would provide phytoremediation with much needed tools. Other targets of biotechnology might include altered lignin production or incorporation into cell walls (for increased phytostabilization), increased lipid concentration or quality in the roots (for all forms of phytoremediation), altered root permeability, and altered infectivity by mycorrhizae.
3. Plant Transformation Unfortunately, not all species of plants are equally amenable to transformation and regeneration. After nearly a decade’s worth of work, monocots and many trees are still proving difficult to transform and regenerate on a routine basis. Since many of the target plants for phytoremediation include these types of plants, we believe continued effort in plant transformation will eventually prove useful to phytoremediation. In one species, now in trials, altered rooting morphology was obtained by Agrobucferium rhizogenes transformation (Han et af., 1993). The resulting trees had greatly increased root mass, surface area, and soil/root contact. This shows that some tree species are indeed amenable to molecular techniques, and schemes for their improvement in phytoremediation have been proposed (Stomp ef al., 1994).
4. Microbial Biotechnology Not all biotech efforts in phytoremediation are directed to the plant component. Efforts at creating “biased rhizospheres” where microbes have additional degradative capacities are ongoing in a number of labs working with biocontrol of plant pests, nitrogen fixation, plant-growth promotion, and myconhizae.
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Few workers have attempted to combine molecular biology, microbial competition in the rhizosphere, and microbial degradation of xenobiotics. One exception is research by Brazil and co-workers (1995). They have inserted the genes (bph) encoding the biphenyl degradative pathway into the chromosome of two rhizosphere pseudomonads. Results of tests on the genetically engineered organism demonstrate that growth rate, bph gene expression and stability, and colonization potential of the rhizosphere were not seriously affected. It may therefore be possible to genetically engineer rhizosphere competent pseudomonads without compromising their competence. Importantly, expression of the bph genes was detected in rhizosphere soil microcosms. The authors suggest that expanding the degradative capabilities of rhizosphere-competent microorganisms might be a good method for generating useful strains for bioremediation applications.
IX. CONCLUSIONS Phytoremediation is an exciting nascent technology. Its development represents an opportunity for cross-discipline research teams to produce a new, and much needed, technology to remediate environmental contamination. Research at all levels is needed. We lack many of the fundamental understandings on the interactions between the system components (plants, their microbial communities, contaminants, and soil, water, and the atmosphere) and field and engineering installation and equipment development. The ability to use natural ecosystems to remediate the environmental damage done by industrial and urban activites has generated excitement among technologists, owners of contaminated sites, regulators, and the popular press. Delivering a widely applicable technology, acceptable to the scientific, regulatory, and political communities, is the current challenge before the research community.
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Allen, M. F. (1992). “Mycorrhizal Functioning: An Integrated Plant-Fungal Process.” Chapman and Hall, New York. Ambler, J. R., Morgan, P. W., and Jordan, W. R. (1992). Amounts of zeatin and zeatin riboside in xylem sap of senescent and nonsenescent sorghum. Crop Sci. 32, 41 1-419. Anderson, T. A , , and Walton, B. T. (1992). “Comparative Plant Uptake and Microbial Degradation of Trichloroethylene in the Rhizospheres of Five Plant Species: Implications for Bioremediation of Contaminated Surface Soils,” p. 186. ORNL/TM-12017. Oak Ridge, TN. Anderson, T. A., Guthrie, E. A., and Walton, B. T. (1993). Bioremediation in the rhizosphere. Environ. Sci. Tech. 27, 2630-2636. Anderson, T. A,, Kruger, E. L., and Coats, J. R. (1994). Enhanced degradation of a mixture of three herbicides in the rhizosphere of a herbicide-tolerant plant. Chemosphere 28. 1551- 1557. Anderson, T. A , , and Coats, J. R . (1995’). Screening rhizosphere soil samples for the ability to mineralize elevated concentrations of atrazine and metolachlor. J. Environ. Sci. Health Part B Pest. Food Contam. Agric. Wastes 30, 473-484. Anderson. T. A.. and Walton, B. T. (1995). Comparative fate of 14C-trichloroethylene in the root zone of plants from a former solvent disposal site. Envimn. Toxicol. Chem. 14, 2041-2047. April, W., and Sims, R. C. (1990). Evaluation of the use of prairie grasses for stimulating polycyclic aromatic hydrocarbon treatment in soil. Chemosphere 20, 253-265. Atlas, R. M., and Bartha, R. (1993). “Microbial Ecology: Fundamentals and Applications.” p. 563. Benjamin/Cummings, Menlo Park, CA. Baker, I. M.,and Van Bavel. C. H. M. (1987). Measurement of mass flow of water in the stems of herbaceous plants. Plant Cell Environ 10, 777-782. Baker, A. J. M., Reeves, R. D., and McGrath, S. P. (1991). In situ decontamination of heavy metal polluted soil using crops of metal accumulating plants-A feasibility study. In “In situ Bioreclamation Applications and Investigations for Hydrocarbon and Contaminated Site Remediation.” (R. L. Hinchee and R. F. Olfenbuttel, Eds., pp. 6061-605. Butterworth-Heinemann, Boston MA. Barber, D. A., and Gunn, K. B. (1974). The effect of mechanical forces on the exudation of organic substances by the roots of cereal plants grown under sterile conditions. New Phytol. 73, 39-45. Bedard, D. L., Unterman, R., Boop, L. H., Brennan, M. J., Haberl, M. L., and Johnson, C. (1986). Rapid assay for screening and characterizing microorganisms for the ability to degrade polychlorinated biphenyls. Appl. Environ. Microbiol. 51, 76 1-768. Bell, R . M. (1992). “Higher Plant Accumulation of Organic Pollutants from Soils.” EPA/600/R-92/ 138, Cincinnati, OH. Bollag, J-M., Myers, C . , Pal, S . , and Huang, P. M . (1995). The role of abiotic and biotic catalysts in the transformation of phenolic compounds. Iti ”Environmental Impacts of Soil Component Interactions” (P. M. Huang. I. Berthelin. J.-M. Bollag. W. B. McGill. and A. L. Page, Eds.). Lewis, Chelsea, MA. Boulter, D.,Jeremy, J. J., and Wilding, M.(1966). Amino acids liberated into the culture medium by pea seedling roots. Plant Soil 24, 121-127. Bowen, G. D. (1969). Nutrient status effects on loss of amides and amino acids from pine roots. PIunt Soil 30, 139- 142. Brazil, G. M., Kenelick. L.. Callanan, M., Haro, A,, DeLorenzo, V., Dowling, D. N., and O’Gara, F. ( I 995). Construction of a rhizosphere pseudomonad with potential to degrade polychlorinated biphenyls and detection of bph gene expression in the rhizosphere. Appl. Environ. Microbiol. 61. 1946-1952. Briggs, G . G . . Bromilow. R. H., and Evans, A. A. (1982). Relationship between lipophilicity and root uptake and translocation of non-ionised chemicals by barley. Pestir. Sci. 13, 495-504. Brcck. T. D. (1966). “Principles of Microbial Ecology.” p. 306. Prentice-Hall, Englewood Cliffs, New Jersey.
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BIOLOGICAL CONTROL OF WEEDS WITH PLANTPATHOGENS AND MICROBIAL PESTICIDES~ David 0.TeBeest Department of Plant Pathology, University of Arkansas, Fayetteville, Arkansas 72701
I. Introduction 11. Strategies for the Control of Weeds with Plant Pathogens A. The Bioherbicide Strategy B. The Classical Strategy C. The Augmentation Strategy 111. Biological Control of Weeds with Plant Pathogens A. Control of Weeds with Microbial Pesticides B. Control of Weeds with the Classical Strategy C. Biological Control of Aquatic Weeds D. Biological Control of Weeds with Bacterial Plant Pathogens IV Biological Control of Weeds by Microbial Management of Seed Banks V. Synergisms That May Affect the Effectiveness of Microbial Agents A. Synergism of Pathogens with Other Pathogens B. Synergism of Pathogens with Insects C. Synerpsm of Pathogens with Chemicals VI.The Environmental Impact of Microbial Herbicides A. Effects on Nontarget Beneficial Plant Species B. Effects on Nontarget Species and Food Web Components VII. Summary References
I. INTRODUCTION The science of using plant pathogens to control weeds is almost as old as the science of plant pathology (Templeton ef al., 1979; Wilson, 1969). Wilson 'Published with the approval of the Director, Arkansas Agricultural Experiment Station, Manu-
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( 1969) reported that Halsted ( 1894) received a letter in 1893 suggesting that if a rust-causing disease of a thistle could be “disseminated through the country, that it (the thistle) would receive a substantial check.” Wilson (1969) further described previous efforts to use pathogens for control of cactus, mistletoe, aquatic and agronomic weeds, and weedy trees that represents a continuous effort in biological control of weeds from 1890 through 1969. Cockayne (1910) reported that fungi had been investigated as “weed controllers” in many parts of the world but without success. Cunningham (1927) reported that “natural control” of weeds with plant pathogens had received “much attention in recent years” for eliminating weeds without direct labor or monetary expense and described modest efforts to control weeds with pathogens in New Zealand. More recently, biological control of weeds with plant pathogens has received renewed interest after the deliberate importation and release of Pucciniu chondrillina Bubak and Syd. in Australia and the United States to control infestations of agricultural, range and forested lands by rush skeletonweed, Chondrillu juncea L. Also, the nearly simultaneous demonstration that commercial formulations of Phytophthora palmivora But1. and Colletotrichum gloeosporiaides (Penz) Penz. & Sacc. could effectively and safely control Morrenia odorata (H. & A.) Lindl. in Florida and northern jointvetch (Aeschynomem virginica (L.) BSP.) in Arkansas clearly demonstrated that plant pathogens could be processed and used as registrable commercial products for weed control with mutual value to commercial and agricultural enterprises. These milestones resulted in the heightened search for organisms such as Colletotrichum, Phytophthora, and Pucciniu with similar usefulness as microbial agents for weed control throughout the world. Many fungi and bacteria subsequently have been evaluated for effectiveness and commercial utility. The successful importation and commercialization of pathogens also has led to the development of several strategies or approaches to utilize pathogens as agents for biological control.
11. STRATEGIES FOR THE CONTROL OF WEEDS WITH PLANT PATHOGENS
A. THEBIOHERBICIDE STRATEGY As bioherbicides (microbial pesticides), pathogens can be applied to control weeds within a specific geographical site (i.e., a single field) by an inundative application of inoculum (Charudattan, 1984). This approach is also referred to as the mycoherbicide approach (Templeton et a l . , 1979). Inundative application of inoculum of pathogens, often to early stages of weed growth, results in the control of weed infestations without the disease developing beyond the initial
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lesions into epidemics. The initial lesions caused by the applied inoculum directly cause the death of infected weed seedlings.
B. THECLASSICAL STRATEGY In the classical strategy, a pathogen is simply released into weed populations (from one to many independent sites) and is expected to increase and disperse naturally throughout the entire weed population without significant subsequent annual releases or augmentation of established populations (Charudattan, 1984). The natural increase of disease on susceptible plants is relied upon to control weeds, either directly from plant death or indirectly through reduction of plant vigor and seed production, over broad geographical settings and within many ecological niches.
C. THEAUGMENTATION STRATEGY The augmentation strategy is similar to the bioherbicide strategy in that while there is direct human manipulation and distribution of inoculum, the inoculum is neither mass-produced nor applied as an inundative dose over large areas (Charudattan, 1984). Control of the weed results from and requires the increase of disease through many disease cycles to reach threshold levels that cause the death of infected plants within treated areas. These strategies permit the utilization of many different types of pathogens causing different types of diseases. While the ability of a pathogen to sporulate on infected tissues and to disperse from infected plants is crucial to the effectiveness of the classical and augmentative strategies, it is not nearly as important when pathogens are utilized as an inundatively applied bioherbicide. For bioherbicides, the ability to infect without environmental or temporal constraints and to cause disease and plant mortality within a short time (i.e., one disease cycle) is more important.
111. BIOLOGICAL CONTROL OF WEEDS WITH PLANT PATHOGENS
A. CONTROL OF WEEDS WITH MICROBIAL PESTICIDES Three endemic fungal plant pathogens have been registered as microbial pesticides in the United States and Canada for control of weeds in agricultural systems. One is a soil-borne fungus, Phytophthoru, while two are foliar pathogens in the genus Collrtotrichum.
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The endemic fungal pathogen Phytophthoru pulmivoru (Feichtenberger er ul., 1984) (= Phytophthora citrophthora ( R . E. Sm. and E. H. Sm.) (Burnett e t a l . , 1973, 1974) was first used commercially as DeVine in 1981 to control stranglervine (=Milkweed vine), Morreniu odoruta, in citrus groves (Kenney, 1986). Morrenia odorutu was introduced from South America and has become a serious weed problem in citrus, competing for water, sunlight, and nutrients, girdling tree limbs, and interfering with harvesting, pesticide applications, and irrigation. The weed infests approximately 120,000hectares in Florida. Phytophthoru palmivoru was initially isolated in 1972 from diseased and dead plants found in Orange County, Florida. In small-scale field tests, 96% of the vines were killed within 10 weeks after infestation of soil with the fungus (Burnett et at., 1973; Ridings et al., 1976, 1977). Colletotrichum gloeosporioides fsp. aeschynomene was developed in the United States and marketed as a microbial pesticide (mycoherbicide) in 1982 as COLLEGO for the control of northern jointvetch (NJV), A . virginica, in rice and soybeans in several states in the lower Mississippi River delta (Bowers, 1986). Northern jointvetch reduces the quality of milled rice and at densities of 1 to I 1 plants/m2, also reduces grain yields from 4 to 19%, respectively. In 1980, 11% of the rice crop was discounted due to the presence of weed seeds (Smith et al., 1973), and the discount loss in Arkansas was estimated to be $7.6 million. Viable spores of the fungus were formulated as a wettable powder to be used, handled, and applied much like any postemergence chemical herbicide. Collerotrichum gloeosporioides f. sp. aeschynomene causes an anthracnose on NJV seedlings, infecting stems, petioles, and leaflets (Daniel et ul., 1973; TeBeest, 1982, 1988, 1990; TeBeest et al., 1978a,b). Enlargement and coalescence of stem lesions result in the girdling and death of the plant above the lesions. The fungus sporulates profusely on the lesion surfaces, and rainfall contributes to dispersal of the fungus spores on the plant, increasing the severity of infection. The fungus also is dispersed by infected seed (TeBeest and Brumley, 1978) and by rain-splash (Yang and TeBeest, 1992a). Small-scale field tests demonstrated that 100% of the seedlings inoculated with the fungus were controlled within a few weeks after treatment (Templeton et ul., 1981). In the hands of growers, the commercial formulation of the fungus provides greater than 90% control of NJV when used according to label directions (Bowers, 1986). COLLEGO has not been marketed by the registrant since 1992. A third mycoherbicide, BIOMAL, is composed of spores of C. gloeosporioides (Penz.) Sacc. f.sp. malvue and was registered in 1992 in Canada for the control of round-leaved mallow (Mulvupusilla Sm.) in wheat (Triticum aestivum L.) (Grant et al., 1988, 1990; Mortenson, 1988, 1991). Round-leaved mallow plants inoculated with spore suspensions were killed within 17 to 20 days after inoculation. The fungus infects leaves, petioles, stems, and crowns of this weed and kills the plant within a few weeks after application. Control is 90 to 100% effective in the field and the infected plants do not reappear the following year.
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The fungus infects several Malvu species, velvetleaf (Abutilon theophrasti Medic.), and hollyhock (Althea rosea (L.) Cav.), but the disease is severe only on M . pusilla. Though registered, BIOMAL has not been available commercially. The following examples of other fungal pathogens under investigation illustrate the variety of targeted weeds and organisms that are being evaluated as biological control agents. Colletotrichum orbiculare (Berk. et Mont.) v. Arx is being reevaluated as a biological control agent for Bathurst bur (Xanthium spinosum L.) in Australia (Auld et al., 1990; Auld and Tisdell, 1985) after tests conducted earlier were occasionally promising (Butler, I95 I). When applied as a mycoherbicide, the fungus controlled 50 to 100% of the seedlings in field tests conducted in 1987 and 1988. The highest levels of control, 98 to loo%, were achieved in a dryland grazing site (Auld et a / . , 1990). The efficacy of the fungus increases with increasing periods of high humidity, and the presence of the extracellular conidial matrix hastened the onset of disease symptoms and increased disease levels on X . spinosum (McRae and Stevens, 1990). In Japan, two fungi, Drechslera rnonoceras and Epicoccosorus nematosporus are being investigated for control of two of the four major weeds in rice fields in Japan. Drechslera monoceras has been reported to give excellent control of barnyardgrass, Echinochloa species, in greenhouse and field tests (Gohbara and Yamaguchi, 1993). Combined use of this fungus and the herbicide pyrszosulfuronethyl controlled most of the weeds growing in paddy fields. Similarly, E . nematosporus has been repeatedly effective in controlling water chestnut (Elocharis kuroguwai) in greenhouse and field tests (Gohbara and Yamaguchi, 1993; Suzuki, 1991). Spore suspensions sprayed onto plants in June and July decreased plant populations and the number of tubers from which plants emerged into the next year. Despite the success of DeVine, much of the work in microbial pesticides has focused on microbial control of weeds by postemergence application of plant pathogens to foliage. However, Boyette er al. (1984), Weidemann (1988), and Weidemann and Templeton (1988) have demonstrated that the soilborne fungus Fusarium solani (Mart.) Appel & Wr. f.sp. cucurbirae effectively controls Texas gourd (Cucurbita texana A. Gray) following preemergence application of inoculum to soil. Inoculum consisting of either spores or alginate granules containing spores and mycelium controlled up to 95% of emerging seedlings. The soilborne fungus Sclerotinia sclerotiorum (Lib.) de Bary also has been investigated for control of Canada thistle (Cirsium arvense (L.) Scop.), spotted knapweed (Centuuria maculosa Lam.), and dandelion (Turuxucumoficinale Weber) (Brosten and Sands, 1986; Miller et a l . , 1989; Riddle et al., 1991). Populations of dandelions in turfgrass were reduced 80 to 85% following repeated applications of heat-killed perennial ryegrass seed infested with S. sclerotiorum (Riddle et a/., 1991). Kentucky bluegrass (Pou pratensis L.), creeping bentgrass (Agrostis pulustris L.), annual bluegrass (Poa annua L, ), and quackgrass
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(Agropyron repens (L.) Beauv.) were not infected. The potential of S. sclerotiorum as a biological control agent for thistle was demonstrated in the United States in Montana (Brosten and Sands, 1986). In field trials, 20 to 80% of Canada thistle shoots were killed after treatment with the fungus applied as sclerotia or as infested wheat kernels, and plant populations were also reduced the following year. The fungus also has been reported to be a pathogen of Centaurea difusa (Watson et a!. , 1974). Sclerotinia sclerotiorum has a very wide host range and infects many plant species of economic value. A genetic approach has been utilized to make the fungus environmentally safe by reducing the effective host range of this pathogen. Miller et al. (1989) reported that an auxotrophic strain was avirulent to four of seven susceptible hosts unless an exogenous source of cytosine was applied at the inoculation site. The apparent virulence of the auxotrophic strain was dependent on the inoculum used in these tests. Inoculum consisting of infested millet seed with no amendments added prior to inoculation resulted in infection of lettuce (Lactuca sativa L.), clover (Trifolium hybridum L.), and sunflower (Helianthus annuus L.), whereas inoculation with PDA cultures without additives resulted in infection of lettuce. The addition of yeast extract or cytosine increased the number of plants of the seven test species infected. Such a genetic approach to limiting host range may permit the use of pathogens with broad host ranges, such as S. sclerotiorum, as bioherbicides (Miller et a l . , 1989). In the United States an endemic rust has also been evaluated for control of a weed utilizing the augmentative approach (Bruckart and Hasan, 1991) rather than a truly classical approach. Pucrinia canaliculatu (Schw). Lagerh. has been evaluated for control of nutsedges, Cyperus rotundus L. and C . esculentus L., in the United States (Callaway et al., 1985; Phatak et a l . , 1983). When released early in the spring, the rust inhibits flowering and tuber formation (Calaway et a!. , 1985). Phatak er al. (1983) had earlier reported that a simple release of the rust into plots resulted in reductions in root growth and fresh and dry weight of shoots. This fungus is capable of rapid dispersal and infection. Within 60 days, approximately 78% of the leaf area was rusted in these tests. In one test, rust pustules were found in a previously healthy stand of yellow nutsedge within 12 days after a single pot of rusted seedlings were placed in similar yellow nutsedge stand 7 km away. An epiphytotic reportedly developed over the entire area. The rust appears to be adapted to a wide range of environmental conditions since epiphytotics developed over several hectares following releases made thoughout the growing season under a variety of conditions.
B. CONTROL OF WEEDS WITH
THE
CLASSICAL STRAGEGY
Pathogens used in the classical approach are expected to reduce weed populations to economically insignificant levels as a result of the natural epidemics they would cause.
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The introduction of the rust fungus Puccinia chondrillina into Australia in 1971 from the Mediterranean region for the control of rush skeletonweed (Chondrilla juncea) appears to constitute the first deliberate introduction of a pathogen for weed control in any country in what has become known as the classical approach to biological control of weeds with plant pathogens (Cullen et a ( . , 1972). Factors that contributed to excellent control of the weed in Australia included rapid dispersal of spores, a high density of susceptible plants (C. juncea), virulence of the pathogen to the dominant biotype of the weed, and favorable weather conditions (Hasan, 1972; Hasan and Jenkins, 1972; Hasan and Wapshere, 1973). Two strains of P . chondrillina from Eboli, Italy, were introduced into the United States in 1975 (Adams and Line, 1984; Emge e t a l . , 1981). Within 2 years, the fungus caused severe infections of plants throughout populations of skeletonweed in California, Oregon, Idaho, and Washington (Lee, 1986; Supkoff et al., 1988). In 1975, Enryloma agerutinae sp. nov. (Barreto and Evans, 1988) was introduced into Hawaii from Jamaica to control hamakua pamakani (Ageratina riparia (Regel) K.&R.) (Trujillo, 1976, 1985; Trujillo er al., 1988). Ageratina riparia was determined to be the most serious weed pest of the Hawaiian range from 800 to 6500 ft in elevation. Releases of the fungus from November of 1975 through May of 1976 resulted in an epidemic and devastation of hamakua pamakani. Weed populations were reduced from 80 to <5% of the original population within 1 year. Control of the weed was most effective between 1500 and 6000 ft elevation where the temperature and rainfall were conducive of disease development. More than 50,000 hectares of pastureland have been rehabilitated by removal of the weed by C. ageratinae. Coiletotrichum gloeosporioides f . sp. clidemiae was isolated from diseased leaves of Clidemia hirta (L.) D. Don collected in Panama by Trujillo et al. (1986) and Trujillo (1992). Clidemia hirta is a major noxious weed of tropical American origin and was introduced accidentally into Hawaii in 1941 where it now infests more than 40,000 ha of rain forest areas on Oahu. It has also spread to several other Hawaiian islands. Trujillo er al. (1986) have found that C. gloeosporioides f.sp. clidemiae is an aggressive pathogen and specific to C. hirta and concluded that this pathogen can be an effective biological control agent for this weed in Hawaii if introduced. A study of Septoria passijlorae in Hawaii by Trujillo er al. (1994) confirmed that this pathogen is specific to Passzjlora rripurtira (Juss.) Poir. var. tripartita Holm-Nie. Jorg. & Laws. This plant was first introduced into Hawaii in 1921 as an ornamental but has become an aggressive, weedy species of high-elevation areas on Kauai and Hawaii. Cultures of S.passiflorue were obtained from infected plants of PassiJora planted near Aldana, Narino, Colombia. Septoriu passiflorue appears to be specific to P . tripartita and P . foetida after host range tests were conducted in Hawaii. Both are introduced weedy species in Hawaii. The fungus does not appear to be pathogenic to economic plants. Since the Pas-
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sifloraceae are not native to Hawaii, S. pussiflorue is presumed to be an environmentally safe biological control agent for these weeds in Hawaii following introduction (Trujillo et uf., 1994). Introduction has been recommended because the probability for success appears to be excellent and host range is very restricted (Trujillo et af., 1994). Currently three imported rust fungi are being evaluated in the United States for biological control of leafy spurge (Euphorbiu esulu L.), musk thistle (Curduus nutuns L. ssp. leiphyflus (Petrovic) Stoj. & Stef.), and yellow starthistle (Centuureu sofstitiafisL.) in the United States (Bruckart and Dowler, 1986; Melching et ul., 1983). A study of I8 isolates of Mefumpsoru euphorbiue (Schub.) Cast. collected in Austria, Hungary, Romania, Switzerland and Yugoslavia resulted in only one compatible combination on two collections of cypress spurge (Euphorbiu cyparissius L.) (Bruckart and Dowler, 1986). Several other isolates produced infections that were difficult to maintain on both leafy and cypress spurge (Turner et a / . , 1983). An isolate of Puccinia curduorum Jacky collected in lbrkey was found to be very aggressive on 23 of 27 collections of musk thistle from Canada, France, and the United States (Bruckart and Dowler, 1986). The rust was pathogenic to 8 of 17 Cirsium species tested but was much less aggressive on species other than C. nutans L. The susceptibility of musk thistle and related composites has also been determined (Politis et a l . , 1984). This rust was released for field studies in the United States in 1987 after it was determined that it was not aggressive on globe artichoke (Cynaru scolymus L.) (Politis et al., 1984). Baudoin ef ul. (1993) have suggested that P . carduorum, particularly in combination with insects, can contribute to the reduction of seed production and control of musk thistle. Pucciniu jaceae Otth. was evaluated for control of yellow starthistle with rust collections from Tbrkey. Isolates of the rust were very aggressive on yellow starthistle but also infected safflower (Carthumus tinctorius L.) (Bruckart and Dowler, 1986). The effects of the rust Pucciniu lagenophorue Cooke on groundsel (Senecio vulgaris L.) have been studied under summer and winter conditions in the United Kingdom (Paul and Ayers, 1986a,b, 1987a,b) after introduction into Great Britain in 1961 (Paul and Ayers, 1987b). Field-grown groundsel, infected as a result of inoculations made in the autumn, showed 70% mortality when measured in the spring whereas mortality was only 40% for plants inoculated during the spring and summer. The higher mortality of plants inoculated in the autumn was attributed to infection of the hypocotyls which usually killed the host plants within 1 to 2 weeks and compromised the ability of surviving seedlings to withstand water stress. Infection of hypocotyls in the summer, though as severe, did not result in significant mortality although infection had very substantial effects on growth and reproduction of populations during the summer. Infection
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resulted in a 50% reduction of vegetative growth, a 15% reduction in plant density, and a 24% reduction in floret production. The fungus does not naturally survive at sufficiently high levels to cause significant mortality. In this case, manipulation of the host-pathogen population relationship over time has a large effect on host survival. In glasshouse and field experiments, rust infections of groundsel decreased the competitive ability of groundsel with lettuce, although lettuce yields were not significantly reduced by rust-infected plants until weed densities exceeded 4000 plants per square meter (Paul and Ayers, 1987a). Rust infections reduced the impact of groundsel on lettuce yield without causing any significant increase in groundsel mortality.
C. BIOLOGICAL CONTROL OF AQUATIC WEEDS Several plant pathogens have been or are currently under investigation for biological control of aquatic weeds such waterhyacinth, water milfoil, duckweeds, alligator weed, waterlettuce, or blue-green algae, in a variety of aquatic environments (Joye, 1990). Pathogens of aquatic weeds that have been tested as microbial pesticides include species of Fusarium and Macrophomina, on hydrilla (Hydrilla verticillatu (L.F) Royle), and species of Acremonium, Colletotrichum, Fusarium, Pythium, and Phytophzhoru for control of eurasian watermilfoil (Myriophylfum spicatum L.), but no promising control agents have been found among these isolates (Joye, 1990). Experiments were conducted with potential commercial formulations of Cercospora rodmanii Conway for control of waterhyacinth (Eichhornia crassipes (Mart) Solmes) (Charudattan et a/., 1985; Conway, 1976a,b). This fungus has been released in South Africa for control of waterhyacinth in the Crocodile river using a classical approach (Morris and Cilliers, 1992). An isolate of Colletotrichum gloeosporioides was also tested as a potential biological control agent for eurasian watermilfoil (Myriophyllum spicatum L.) (Smith et a l . , 1989). Under realistic conditions, the effect of this fungus on watermilfoil was too small to warrant further consideration as a possible biological control agent. Recently, Microleptodiscus rerrestris (Gerdemann) Ostazeski was reported to have considerable impact on the populations of milfoil in Florida tests (Joye, 1990). In recent work, Verma and Charudattan (1993) showed that this fungus was pathogenic to 3 (Hydrilla verticilata, Myriophyllum aquaticum, and Ceratophyllum demersum L.) of 16 aquatic plant species tested. Only on Hydrillu did the fungus cause plant mortality comparable to levels achieved by infection of watermilfoil. The remaining 13 species were not infected. Of 17 terrestrial species tested, seed germination was significantly affected by the fungus although postemergent symptoms of disease developed on seedlings of only 10 species. On 4 of these, Medicago sativu L., Lotus corniculatus L., Trifolium
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hybridum L., and T . repens L., disease developed to affect from 26 to 50% of the host plant tissues. Shearer (1994) reported that application of a formulation of M. terrestris was ineffective in reducing aboveground biomass of eurasian watermilfoil under natural conditions in the field, but poor field performance was attributed more to fungus/formulation problems than to biological, chemical, or physical factors encountered in this test.
D. BIOLOGICAL CONTROL OF WEEDS WITH BACTERIAL PLANTPATHOGENS Although bacterial diseases of weeds have been known and described for many years (Rosen, 1924), until recently, few have been investigated for potential biological control of weed species. Recently, Caesar ( 1994) suggested that strains of Agrobacterium tumefaciens (E. F. Smith & Town) isolated from important rangeland weeds may be effective as biological control agents for their respective hosts. Host ranges of isolates of A . tumefaciens from Russian knapweed and leafy spurge and strains representing biovars I and 2 of A . tumefaciens and A . viris varied greatly with six strains being pathogenic to no more than one species in addition to the original host. Strains from New Mexico were highly pathogenic to diffuse knapweed (Centaurea diffusa L.), causing girdling, stunting, and death of this host. Zhou and Neal (1995) recently compared strains of Xanrhomonas campestris (L. R . Jones et a / . ) pv. poannua as biocontrol agents for annual and perennial subspecies of annual bluegrass ( P . annua L.). Results of controlled growth chamber and field tests showed that two strains of this bacterium were similarly virulent in both tests. In growth chamber tests, annual and perennial subspecies of P . annuu were controlled 82 and 92%.respectively, while in field tests control reached only I 1 and 7%, respectively, following repeated weekly applications. Control of annual bluegrass was only 40% following 4 weeks of three applications per week. The weeds also recovered 2 to 5 weeks after weekly inoculations were stopped. Johnson (1994) reported that three applications of two strains of X . campestris pv. poannnua controlled between 52 and 82% of the annual bluegrass in dormant bermudagrass (Cyanodon transvaafensis Burtt-Davy X C . ducryfon (L) Pers.) field plots. Begonia et al. (1990) have demonstrated in culture tube assemblies that isolates of Pseudomonas and Erwinia herbicola caused velvetleaf (Abutilon theophrasti) seedlings to become chlorotic and develop abnormal root systems compared to noninoculated controls.
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IV. BIOLOGICAL CONTROL OF WEEDS BY MICROBIAL MANAGEMENT OF SEED BANKS As seen above and as widely reported in previous reviews (Templeton et a l ., 1979; TeBeest, 1991), considerable attention has been given to developing pathogens as microbial pesticides or as classical agents for control of weeds. Weed seed banks are considered to be the major source of weed infestations in arable lands (Cavers and Benoit, 1989). Recent work by Kremer (1993) and Kremer et al. (1990) has suggested that deleterious rhizobacteria may be useful in reducing weed seed banks in yet another approach to reducing weed infestations in crop and range lands. Kennedy et al. ( 1991 ) screened more than 1000 isolates of pseudomonads, and 81 inhibited downy brome (Bromus rectorum) but not wheat. Six isolates consistently inhibited downy brome growth but not wheat in controlled environments. In some field tests, isolates reduced downy brome populations up to 30% and shoot dry weights by 42%. Winter wheat yields were increased in two of three field locations because of suppression of downy brome.
V. SYNERGISMS THAT MAY AFFECT THE EFFECTIVENESS OF MICROBIAL, AGENTS The term synergism is used loosely here to mean a combined use of insects, chemicals, or pathogens that enable pathogens to control weeds when the individual activities of the interactive participants are less effective. Synergism as used here should also not be confused with the integrated use of various components (i.e., biological and chemical pesticides) that may or may not be inhibitory of each other’s activities but that, nevertheless, have been employed in effective control schemes employing both components (Klerk et al., 1985).
A. SYNERGISM OF PATHOGENS WITH *HER
PATHOGENS
Several examples have been reported in which pathogens incapable of causing significant levels of disease when infecting alone were more severe in combination with other pathogens. For example, Dimock and Baker (1951) showed that Fusariiim roseum Lk. emend. Snyd. & Hans. infected snapdragon (Antirrhinum majus L . ) through lesions caused by the rust fungus Puccinia antirrhini D.& H. Apparently, F . roseum infected healthy tissue beyond the rust lesion and caused
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death of leaves and shoots or even entire plants whereas infection by the rust alone seldom caused death. Rust-free plants, even when severely wounded, were not infected by F. roseum. Thus it appears that a facultative parasite, incapable of infecting a plant alone, contributed to increased disease severity by invading lesions produced by another pathogen. This phenomenon has been extended to biological control of weeds. Alternaria macrospora Zimm. has been investigated as a potential mycoherbicide to control spurred anoda (Anoda crisrata (L.) Schlecht.) in the United States (Crawley et al., 1985; Walker, 1981; Walker and Sciumbato, 1979). The susceptibility of spurred anoda to A . rnacrospora is highly correlated with plant age. Approximately 100% of seedlings inoculated at the cotyledonary stage were killed by infection, but less than 10% were killed at the three- to four-leaf stage (Walker and Sciumbato, 1979). However, 100% of plants inoculated at the fourto five-leaf stage were killed by the interaction of A . rnacrospora and Fusarium lateritium Nees ex Fr. Fusarium lateritium is a weak pathogen of spurred anoda and causes less than 20% mortality when inoculated to seedlings alone, although F. laterilium usually killed wound-inoculated spurred anoda seedlings (Crawley et al., 1985). The highest mortality occurred when Afternaria was inoculated 5 days before Fusarium. When Fusarium was inoculated 5 days before Alternaria, only 1 1 % of the seedlings were killed. One suggested explanation for the Afternarial Fusarium interaction was that F . lateritium penetrated and infected through the lesions caused by Alternaria. These results indicate that sequential applications of both fungi were more effective than either fungus used alone for control of spurred anoda. Mortality of groundsel infected by P . lagenophorae has been attributed to invasion of rust lesions by Botntis cinerea Pers. (Hallett et al., 1990a,b). Inoculation of healthy groundsel with B . cinerea caused only 10% mortality and only 40% mortality of abiotically wounded plants: however, all plants previously infected by P . fagenophorae died after inoculation with Botrytis. Death of plants was attributed to growth of Botryris into stems. The time necessary to kill plants was dependent upon several factors, including the inoculum concentration of Botrytis and initial pustule numbers of the rust (Hallett, 1990a.b).
B. SYNERGISM OF PATHOGENS WITH INSECTS Insects and plant diseases may have played a role in the control of prickly pear in Australia and other countries (Wilson, 1969). However, the relative importance of pathogens in biological control of prickly pear is not clear since most of the intensive research focused on the role of insects and Cactoblastis cacrorum Berg rather than on potential fungal and bacterial pathogens (Wilson, 1969). Similarly, introduction of the insect Proceidorchures utilis Stone into Australia
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from Hawaii in 1952 may have coincidentally introduced a pathogen, Cercospora eupatorii, that resulted in the build-up of a leaf spot disease of crofton weed, Eupatorium adenophorum Spreng. The combined effect of this insect and pathogen (with native insects) may have contributed to control of crofton weed (Dodd, 1961). Cercospora rodmanii is a pathogen of waterhyacinth, Eichornia crassipes, with biological control potential as first described by Conway (1976a). This fungus causes a leaf spot disease, is highly virulent to water hyacinth and is capable of inflicting severe damage on waterhyacinth (Conway, 1976b; Charudattan, 1986). Waterhyacinths are infested with a variety of fungi and bacteria and at least two insects, the weevil Neochefina eichorniae Warner, and a mite, Orthogaliimna terebrantis Wallwork (Charudattan et al. , 1978). It was generally concluded that arthropod-infested plants were more diseased than plants without insects (Charudattan el al., 1985). Charudattan (1986) concluded that insects and Cercospora in combination could control up to 99% of the waterhyacinth plants in treated areas within 7 months after treatment and that their interaction was synergistic in nature. More recently, Yang and TeBeest (1992b, 1994) showed that green treefrogs (Hyla cinerea Schneider) and grasshoppers, Conocephalis spp. and Melanoplus diTerentialis (Thos.), may be important vectors in the dispersal of inoculum and disease caused by C . gloeosporioides f.sp. aeschynornene on northern jointvetch. Their reports, however, did not attempt to quantify the role these two vector groups played in the control of northern jointvetch by this commercial biological pesticide. It was noted, however, that most of the treefrogs carried the pathogen on their bodies, while only a small but significant fraction of the grasshoppers captured in rice fields carried the pathogen on their mouthparts or feet. A strain of the fungus Colletotrichum gloeosporioides has been associated with mortality of St. John’s-wort, Hypericum perjoratum L., and is under investigation as a potential biological control agent for that weed (Hildebrand and Jensen, 1991) in Canada. An insect, Chrysolina hyperici Forst., has been intoduced and is established in Nova Scotia; it is often observed feeding on St. John’swort but has not given adequate control. Although the fungus effectively spread to control plots in field tests, Hildebrand and Jensen (1991) did not address whether this dispersal resulted from insect or environmental factors or if the disease and insect acted together to help control this weed. It is also possible that even while insects and pathogens may occur on the same host plant, an interaction is not established. For example, Linders (1994) studied the interaction between fungal pathogens and an insect herbivore on Planfago lanceoluta L. After a thorough study, it was shown that a weevil, Trichosirocalus froglodytes Fabr., had only marginal if any interaction with a pathogen, Diaporthe adunctr (Rob.) Niessl., that affected the ability of the pathogen to infect and control this weed.
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C. SYNERGISM OF PATHOGENS WITH CHEMICALS A promising consideration concerning possible chemical interactions with plant pathogens is the utilization of various chemicals, including herbicides, to increase the effectiveness of biological agents by weakening host resistance to infection by a pathogen (Charudattan, 1993). Colletorrichurn coccodes (Walk). Hughes is a pathogen of two nightshade species, Solanurn ptycantlium Dun. and Solanum sarrachoides Sendt., and velvetleaf, Aburilon rheophrasti Medik. (Anderson and Walker, 1985; Eberlein et al., 1991; Wymore et al., 1987, 1988). Although the fungus infected velvetleaf over a wide range of dew period durations and temperatures, disease was most rapid and destructive at 24°C following a 24-hr dew period. Velvetleaf plants at all growth stages were susceptible and were reduced in vigor following inoculations, but only plants at the cotyledonary stage were killed (Wymore et a [ . , 1988). Under less than optimal conditions, the fungus was unable to kill seedlings in the field. Tank mix applications of C . coccodes and thidiazuron ( N phenyl-N'- 1,2,3-thidiazol-5-yl-urea) acted synergistically to increase velvetleaf mortality when compared with the fungus alone (Wymore et a f . , 1987; Hodgson el al., 1988). Gohbara and Yamaguchi (1993) showed that the combined use of the herbicide pyrazosulfuron-ethyl and the fungus D . monoceras showed significant synergisim in controlling barnyardgrass in rice in Japan. Application of pyrazosulfuron-ethyl at 180 mg/a with 109 conidia/a caused 100% mortality of seedlings, exceeding control of seedlings by pyrasulfuron alone or with similar concentrations of conidia alone. Similarly, Scheepens ( 1987) showed that atrazine was synergistic with the pathogen Cochliobolus lunutus for control of Echinochlou crus-galli (L.) Beauv. in maize (Zea mays L.) in The Netherlands. In controlled experiments in a greenhouse or growth chambers, barnyardgrass seedlings could be controlled with the fungus after treatment with a sublethal dose of atrazine. Alternuria cassiae Jurair & Khan was intensively studied as a potential bioherbicide for sicklepod, Cassiae obtusifolia L, in the United States. Sharon et al. (1992) showed that suppression of a phytoalexin biosynthesis with a sublethal dose of glyphosate increased the susceptibility of C . obtusifolia to infection by A . cassiue. Also, five times less inaculum was needed to control sicklepod when applied with glyphosate than without the herbicide, without affecting its specificity. This technique offers an intriguing approach to enhance the pathogenic effects of many pathogens and, therefore, also to increase their potential effectiveness as microbial pesticides. In yet another unique approach to increase the effectiveness of biological pesticides by aiding the establishment of infections in the abscence of conducive environmental conditions, Boyette et al. (1993) and Egley er a f . (1993) used invert emulsions to reduce the free moisture requirements for germination of
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conidia of Colletotrichum truncatum (Schw.) Andrus & W. 0. Moore to control hemp sesbania, Sesbania exaltata L. When sesbania seedlings were sprayed with aqueous suspensions of conidia and were not given a postinoculation dew treatment, only a few seedlings were killed by infection. However, 100% of the seedlings sprayed with invert emulsions of condia that did not receive a dew treatment were killed within 7 days after inoculation. These results were as expected because the moisture requirements for conidial germination were satisfied by the invert emulsion. Similarly, Amsellem et al. (1990, 1991) showed that the inoculum thresholds needed to control C . obtusifolia and Datura stramonium L. with A . cassiae and Alternariu crassa (Sacc.) Rands, respectively, were reduced to one conidium per droplet when an invert emulsion was used as an adjuvant for the inoculum. However, invert emulsions also have the potential for crop damage, and the significance of potential crop damage by mycoherbicides in invert emulsions is not clear (Amsellem, 1991). A concern with the use of C . gloeosporioides f.sp. aeschynomene, and potentially for other mycoherbicides on other crops as well, is integration with fungicides used to control rice and soybean diseases. Proper timing of fungicide treatments with C . gloeosporioides f.sp. aeschynomene enabled the effective use of both pesticide groups (Klerk et al., 1985). Colletotrichum gloeosporioides f.sp. aeschynomene has been integrated into the rice management system in Arkansas, which utilizes chemical pesticides to control weeds, insects, and diseases (Kenney, 1986; Klerk et al. 1985).
VI. THE ENVIRONMENTAL IMPACT OF MICROBIAL HERBICIDES A. EFFECTS ON NONTARGET BENEFICIAL PLANTSPECIES The ability of plant pathogens of weeds to infect cultivated and noncultivated plant species is a serious and important part of every effort to develop a commercial product for biological control of weeds. Almost without exception, these pathogens can and do infect cultivated and horticulturally important plant species in controlled experiments. And of course, numerous pathogens of crop plants also infect weeds. Because of necessity and importance, various schemes to identify the susceptible plant species and the host range of each plant pathogen under consideration for weed control have been developed (Wapshere, 1974). Colletotrichum gloeosporioides f. sp. aeschynomene has been used in a registered product, COLLEGO, since 1982 to control a single weed species, A . virginica, in rice, Oryza saliva L., and soybeans, Glycine max Merr., in the
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United States. This fungal pathogen infects several species within the genus Aeschynomene. However, it also infects species within eight other genera, including Cicer, Indigofera, Lathyrus, Lens, Lotus, Lupinus, Vicia, and Pisum. It is important to note, however, that the fungus is highly virulent only to A . virginica and Lathyrus arboreus L., although certain cultivars of Pisurn sativurn L. were also severely infected (TeBeest, 1988; Weidemann et a l . , 1988). Similarly, the host range of P . palmivora, the fungus formulated as DeVine, includes species other than M. odoruta, the intended target (Ridings et a l . , 1976). The host range of P . palmivora includes onions, cantaloupe, watermelon, okra, tomato, endive, cucumber, english pea, and carrot, and even certain citrus root stocks based on greenhouse tests of preemergence, postemergence, and foliage inoculations. On the other hand, the host specificity of C. gloeosporioides f.sp. malvae appears to be more limited than either of the other two commerically available bioherbicides. Colletotrichum gloeosporioides f. sp. malvae appears to be specific to plants within a single family, the Malvaceae, except for two species, safflower (Carthamus tinctorius) and white mustard (Brassica hirta) (Mortenson, 1988). However, infections of either of these two species did not affect the plants. The fungus infected weedy and ornamental species within the Malvaceae with disease being most severe on certain Malva species and Abutilon theophrasti Medic, which is also a weed. These three examples are sufficient to illustrate two obvious facts. First, weed pathogens can and do infect nontarget cultivated and noncultivated species. Second, infections of nontarget species are occasionally severe. Host range testing protocols, however, have been gradually developed and adopted to identify potential nontarget cultivated and noncultivated host species. Similarly, nearly all host-range tests have been conducted in greenhouses or growth chambers under carefully controlled conditions. It is not certain if host ranges tests conducted and determined in this manner have any direct relationship to similar tests conducted in the field.
B. EFFECTS ON NONTARGET SPECIES AND FOOD WEBCOMPONENTS Colletotrichum gloeosporioides f. sp. aeschynomene has been used in a variety of test systems to evaluate the infectivity of microbial pest control agents on nontarget species (Fournie et a l . , 1988; Genthner and Foss, 1991, 1994). Fournie et al. (1988) showed that this plant parasitic fungal species did not cause adverse effects to a series of test species including fish (Cyprinodon variegatus), a crustacean (Paleomonetes pugio), or a bivalve mollusc (Crassostrea virginica). The test system employed in this work could be used effectively to evaluate other
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microbial agents of different classes and families. For example, Genthner and Foss (1994) showed that Colletotrichum did not cause any adverse effects to Paleomonetes embryos exposed to viable conidiospores. However, Metarhizium anisopliae recently registered by the EPA for control of nuisance flies and cockroaches adversely affected Puleomonetes in comparative tests. Metarhiziurn adversely affected grass shrimp embryos, in some tests killing 67% of the embryos; in some tests, delayed hatch and larval death were also observed. In additional testing, Genthner and Middaugh (1994) noted that toxicity of Metarhizium embryos of inland silverside fish, Menidia beryllina, was probably not due to an artifact of the test protocol but to the organism, since Colletotrichum caused no reaction in parallel studies. A strain of Pseuclomonasfiuorescens also had no effect on nontarget species in these tests (Genthner et u l . , 1991) while potential adverse effects were noted for the insect pathogens Metarrhizium anisopliae, Bacillus sphaericus, and Beauvaria bassiuna (Genthner and Foss, 1991, 1994).
VII. SUMMARY Biological control of weeds with plant pathogens has been an area of study for approximately 100 years. The earliest reports generally illustrated unusual occurrences of disease of weedy species or attempts to control weeds by augmenting populations of pathogens in localized infestations of weeds, often without repeatable success. The more recent and much more successful importations of several pathogens into several countries have served to increase interest in the classical approach to biological control of weeds to such a point that several countries are now actively pursuing this approach. In addition, there is a renewed interest in biological pesticides, and Rodgers (1993) has estimated that biological pesticide sales are increasing by 10-25% per year while agrochemical markets are either static or shrinking. The development of biological herbicides for weeds in the early 1980s has showed that plant pathogens can also be effectively used as biological pesticides. This approach has been developed along with newly implemented guidelines and regulations for their use. Most importantly, they have been effectively and reliably used on a commercial scale. The work on biological control of weeds with plant pathogens has led to a clearer understanding of diseases on noncultivated plant species and on how these pathogens interact with populations of these plant species. Continued investigation of these diseases also may serve to help us better understand diseases of cultivated crops.
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It is nearly certain that additional pathogens will be found, but it is much less certain that they will of economic benefit to users or manufacturers. The economic and ecological benefit of each new potential biological control agent is subject to the complexity of cost/benefit analysis and must consider an enormous number of economic and biological variables (Tisdell, 1987). The enormity and cost of this task will require very careful evaluation by research and industry of potential target weeds and prospective pathogens that can even be considered for introduction or as potential bioherbicides.
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Supkoff, D. M.,Joley, D. B., and Marois, J. J. (1988). Effect of introduced biological control organisms on the density of Chondtilla juncea in California. 1. Appl. Ecol. 25, 1089-1095. Suzuki, H. (1991). Biological control of a paddy weed, water chestnut, with a fungal pathogen. In “Proceedings, International Seminar on Biological Control of Plant Diseases and Virus Vectors” (H. Komada, K . Kiritani, and J. Bay-Petersen, Eds.), pp. 78-86. Food and Fertilizer Technology Center (ASPAC), Taipei, Republic of China on Taiwan. TeBeest, D. 0. (1982). Survival of Collerotrichum gloeosporioides f.sp. ueschynomene in rice irrigation water and soil. Plant Dis.66, 469-472. TeBeest, D. 0. (1988). Additions to the host range of Colterorrichurn gloeosporioides f.sp. aeschynornene. Plant Dis. 72, 16-18. TeBeest, D. 0. (1990). Ecology and epidemiology of fungal plant pathogens studied as biological control agents of weeds. I n “Microbial Control of Weeds” (D. 0. TeBeest, Ed.), pp. 97-1 14. Chapman and Hall, New York. TeBeest, D. 0. (1991). “Microbial Control of Weeds.” Chapman and Hall, New York. and Brumley, J. M. (1978). Colletotrichurn gloeosporioides f.sp. aeschynomene TeBeest, D. 0.. borne within the seed of Aeschynomene virginica. Plant Dis. Reptr. 62, 675-678. TeBeest, D. 0.. Templeton, G.E., and Smith, R. I., Jr. (1978a). Temperature and moisture requirements for development of anthracnose on northern jointvetch. Phvropurhology 68, 389-393. TeBeest, D. O., Templeton, G . E., and Smith, R. J., Jr. (1978b). Histopathology of Colketorrichum gloeosporioides f.sp. aeschynomene on northern jointvetch. Phytopathology 68, 127 1- 1275. Templeton, C . E., Smith, R. I., Jr., TeBeest, D. O., Beasley, J. N. and Klerk, R. A. (1981). Field evaluation of dried fungus spores for biocontrol of curly indigo in rice and soybeans. Arkansas Furm Res. 30(6), 8. Templeton, G . E., TeBeest, D. 0.. and Smith, R. J., Jr. (1979). Biological weed control with mycoherbicides. Ann. Rev. Phyroparhol. 17, 301-3 10. Tisdell, C. (1987). Economic evaluation of biological weed control. Plant Prot. Q.2(1), 10-1 I . Trujillo, E. E. (1976). Biological control of hamakua pamakani with pathogens. Proc. Am. Phytopathol. Soc. 3 , 298. Trujillo, E. E. (1985). Biological control of hamakua pamakani with Cercosporella sp. in Hawaii. In “Proceedings, V1 International Symposium on the Biological Control of Weeds” (E. S. Delfosse, Ed.),pp. 661-671. Vancouver, Canada. Trujillo, E. E. ( 1986). Colletotrichum gloeosporioides, a possible biological control agent for Clidemia hirtu in Hawaiian forests. Pkunt Dis. 70, 974-976.
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Trujillo, E. E. (1992). Bioherbicides. In “Frontiers in Mycology” ( G . F. Leatham. Ed), pp. 196-21 I . Chapman and Hall, New York. Trujillo. E. E., Aragaki, M., and Shoemaker, R . A. (1988). Infection, disease development, and axenic culture of Entylomu cornpositarum. the cause of hamakua pamakani blight in Hawaii. Planr Dis. 72, 355-357. Trujillo, E. E., Norman. D. J.. and Kilgore, E. M.(1994). Septoria leaf spot, a potential biological control for banana poka vine in forests of Hawaii. Plant Dis. 78, 883-885. Turner, S . K . . Bruckart, W. L.. and Fay, P. K . (19x3). European rust fungi pathogenic to collections of leafy spurge from the United States. PAptoputhologv 73, 969. [Abstract] Verma. U. and Charudattan, R . (1993). Host range of Mycoleptodiscus terrestris, a microbial herbicide candidate for Eurasian watermilfoil, M.yririphv//umspicatutn. Biol. Conrrvl 3, 271~
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Walker, H. L. (19x1). Factors affecting biological control of spurred anoda (Anoda cristata) with Alrernuria macrosporu. Weed Sci. 29, 505-507. Walker, H. L., and Sciunibato. G. L. (1979). Evaluation o f Alt~rnarirrmacrosporu as a potential biocontrol agent for spurred anoda (Anodu crisrata): Host range studies. Weed Sci. 27.61 2-614. Wapshere, A. J. (1974). A strategy for evaluating the safety oforganisms for biological weed control. Ann. Appl. Biol. 77, 201-21 I . Watson. A. K . , Copernan. R . J., and Renney, A . J. (1974). A first record of Sclerofinia scleroriorum and Microsphaeropsis cmtaureae on Centaurea diffusa. Can. J . Bofany 52, 2639-2640. Weidemann, G . J. ( 1988). Effects of nutritional amendments on conidial production of Fusarium solmi f.sp. cucurbitue on sodium alginate granules and on control of Texas gourd. Plant Dis. 72. 757-759. Weideniann, G . J., TeBeest, D. 0.. and Cartwright. R . D. (1988). Host specificity of Collerorrichurn gloeosporioides f . sp. aeschvnornetie and C . truncatiim in the Leguminosae. Phyroppurhologv 72, 986-990.
Weidemann. G. J.. and Templeton, G. E. (1988). Control of Texas gourd. Cucurbitu tcxana, with Fiisuri14msolani f.sp. cucurhirae. Weed Technol. 2. 27 1-274. Wilson, C. L. (1969). Use of plant pathogens in weed control. Annu. Rev. Phytopathol. 9,411-434. Wymore, L. A., Poirier, C., Watson, A. K.. et a / . (1988). Colletotrichum coccodes, a potential bioherbicide for control of velvetleaf (Ahirrilon theophrasti). Plant Dis. 72, 534-538. Wymore, L. A , , Watson, A. K., and Gotlieb. A. R. (19x7). Interaction between CoNcfotrirhum coccodes and thidiazuron for control of velvetleaf (Abufilon theophrasfi). Weed Sci. 35, 377383.
Yang, X. B., and TeBeest, D. 0. (1992a). Rain dispersal of Colletotrichum gloeosporioides under simulated rice field conditions. Phvlopathologv 81, I 2 19- 1222. Yang, X. B., and TeBeest, D. 0. (1992b). Green treefrogs as vectors of Col/etotrichum gloeosporioides. Pkutit Dis. 76, 1266- 1269. Yang, X. B.. and TeBeest, D. 0. (1994). Distribution and grasshopper transmission of northern jointvetch anthracnose in rice. Plant Dis. 78, 130-133. Zhou, T., and Neal, J. C. (19%). Annual bluegrass (Poo untwa) control with Xanthomonus cumpestris pv. poannua in New York State. Weed Technol. 9, 173-177.
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ORGANIC AMENDMENTSAND PHOSPHORUS SORPTIONBY SOILS F. Iyamuremye and R. P. Dick Department of Crop and Soil Science, Oregon State University, Corvallis, Oregon 9733 1
I. Introduction A. Phosphorus Cycle in Soils B. Early History 11. Aerobic Soils: Organic Acids and Phosphorus Sorption A. Organic Acids in Soils B. Complexation Reactions with Metals C. Competition for Sorption Sites D. Dissolution of Precipitated Phosphate and Formation of Solid Phosphate Phases E. F,ffects on Surface Charge F. Phytoavailability of Phosphate 111. Aerobic Soils: Plant Residues and h m a l Manures A. Soil pH B. Exchangeable Aluminum and Iron C. Phosphorus Content o f Organic Residues and Phosphorus Sorption D. Biological Transformations of Phosphorus and Fate of Phosphorus from Organic Amendments E. Phosphorus Sorption F. Organic .Amendments Enriched with Inorganic Phosphorus and Phytoavailability of Phosphorus rV. Waterlogged Soils A. Organic Amendments and Eh B. Organic Amendments and pH C. Flooded Soils and Phosphorus Solubility D. Effects of Organic Amendments on Phosphorus Sorption V. Research Needs References
I. INTRODUCTION Phosphorus (P) is an essential nutrient for plant growth and development. It is a deficient nutrient in most soils, particularly in soils with andic properties and
140
F. IYAMUREMYE AND R. P. DICK
highly weathered soils such as Ultisols and Oxisols. Generally, P is available to plants in very small amounts in acid soils, due to adsorption by Fe or Al oxides or by its precipitation with soluble A1 and Fe in acid soils, whereas in alkaline soils phosphate readily reacts with Ca to form insoluble precipitates. Liming soils is the traditional method used to reduce P sorption to increase its plant availability in acid soils. However, liming is an expensive input and is less effective on high P-fixing soils (Haynes, 1982). With the current interest in reduced use of purchased inputs and efficient use of organic residues in agroecosystems, there has been renewed interest in the past 10 years on use of organic amendments to increase P efficiency and availability to plants. This idea has a long history, but recent developments in research and analytical methods have provided significant advances in our understanding of the role of organic amendments in affecting P sorption in soils. Besides addressing organic amendments, the scope of this chapter includes specific components (e.g., organic acids, transformation of organic P in soils) that are important in the relationship between organic amendments and P sorption. Because of divergent chemical and biological processes that occur between aerobic and waterlogged soils, this review is subdivided to separately address these two types of soil environments. More attention is devoted to aerobic soils because considerably more research has been done on the effect of organic residues on P sorption in aerobic than waterlogged soils.
A. PHOSPHORUS CYCLE IN SOILS Phosphorus reactions in the soil environment have been extensively reviewed elsewhere (Wild, 1949; Larsen, 1967; Haynes, 1982; Berkheiser ef al., 1980; Sanchez and Uehara, 1980; Sanyal and De Datta, 1991). Therefore, we will provide a brief overview of P pools and transformations as a framework for understanding the P dynamics when organic residues are added to soils. The P cycle can be characterized as the flow of P between plants, animals, microorganisms, and solid phases of the soil. Major P processes for soils shown in Fig. 1 include P uptake by plants; biological mediated turnover of P through mineralization/immobilizationreactions; and chemical fixation/dissolution reactions between liquid and solid phases. When plant productivity is emphasized, P often has been partitioned into pools based on their potential to provide inorganic orthophosphate for plant uptake. These P pools are broadly envisioned as soil solution P, labile P, and nonlabile P. The labile P is defined as a P reserve that can replenish soil solution orthophosphate in response to P uptake by plant roots. Conversely, nonlabile P has minimal or nonexistent effects on soil solution orthophosphate on an annual basis. Both labile and nonlabile P pools may contain inorganic and organic constituents.
ORGANIC AMENDMENTS AND PHOSPHORUS SORPTION
PLANT ROOTS
141
PLANT RESIDUES
Agure 1 The P cycle in soils, showing the partition of organic and inorganic forms of P into pools based on availability to plants (adapted from Stewart, 1980).
Maintenance of adequate orthophosphate in soil solution for plant growth is controlled by both chemical- and biological-mediated reactions. Ghoshal ( 1975) studied the competition between these two mechanisms for controlling soil sohtion orthophosphate and concluded that, generally, chemical reactions dominate in soils. However, the microorganisms play a critical role in mineralization of organic inputs and partitioning P into various organic fractions and likely provide continuous inputs of orthophosphate into soil solution during the growing season. From this brief overview, it is obvious that organic amendments would have impact on nearly all aspects of the P cycle in soils. The addition of such residues will provide a source of C and stimulate biological activity. Furthermore, the resulting residue decomposition will release organic ligands or acids that will affect P sorption and solubility reactions. Organic amendments also can be a source of inorganic and organic P with the latter being subject to mineralization which releases inorganic P or forms organic P fractions.
1. Chemical Reactions of Phosphorus in Soils Phosphorus sorption is one of the most widely studied reactions in soils. Sorption may include adsorption and precipitation reactions. Sposito (1986)
F. IYAMUREMYE AND R. P. DICK
142
defined adsorption as bidimensional and precipitation as tridimensional. However, he recognized that the two mechanisms are difficult to distinguish, and both mechanisms are described by the same mathematical model. Therefore, for the purposes of this paper, P sorption is defined as the loss of orthophosphate to the solid phases of soils which can occur by either adsorption or precipitation. Bohn et al. (1979) described two adsorption reactions. The first is specific adsorption or ligand exchange where a phosphate anion replaces the hydroxyl on the crystal of hydrous A1 or Fe structures. This mechanism is kinetically described as very rapid and completed within a few days. The second mechanism, nonspecific adsorption, is mediated through the protonation of the hydroxyl surface that creates positive charges and attracts negative charged anions such as orthophosphate. Goldberg and Sposito (1984) have visualized the surface adsorption models as follows:
+ H,PO, S-H2P0, + H 2 0 S-OH + HZPO, * S-HPO, + HZO S-OH + HPO, i3 S-PO, + H,O ,
S-OH
(1) (2) (3)
where S refers to the surface of the solid phase. Other anions such as fluorine, sulfate, and organic acids may be sorbed at the surfaces of A1 or Fe crystals (Stumm et al., 1980; Violante and Gianfreda, 1993; Sibanda and Young, 1986; Parfitt, 1978) similarly to orthophosphate adsorption. These anions then may be in competition with orthophosphate for sorption sites. The precipitation reaction process can be slow and may take years for completion to occur (Bohn et al. 1979; Syers, 1971). As shown in Fig. 2, pH has a major role in P precipitation. Under alkaline conditions, CaZ+ controls P solubility where orthophosphate readily forms less soluble di- and tricalcium phosphates. Under acid conditions, AP+ and Fe3- control solubility of P with orthophosphate readily precipitating as highly insoluble phosphate compounds by the following reactions:
+ PO,3- + 2H,O AlPO, 2H20,,, (amorphous variscite) Fe3+ + Po,’- + 2H,O a FePO, 2H20(,, (strengite) .
A13+
*
*
(4)
(5)
2. Biological Transformation of P and Organic P Fractions Phosphorus is an important element in all biological systems, participating in most metabolic pathways and as a structural component of nucleic acids, coenzymes, phosphoproteins, and phospholipids (Tate, 1984). Biological cycling of P
ORGANIC AMENDMENTS AND PHOSPHORUS SORPTION
143
PH Figure 2 Stability diagram for selected discrete phosphate compounds in aqueous environments (adapted from Berkheiser rt d..1979).
in ecosystems is driven by the presence of plants, soil fauna, and microorganisms. Soil biological activity controls the turnover of P by decomposition of organic P inputs to soils which releases inorganic phosphate or forms organic P fractions, immobilization of P in viable cellular biomass, and solubilization of insoluble mineral P forms through release of chelating agents and/or organic acids. The mineralization/immobilization processes occur simultaneously and net rates are difficult to measure because the end product of mineralization, orthophosphate, is readily subject to fixation reactions in soils. Soil organic P generally makes up 20-80% of total soil P (Stevenson, 1986), and of the P in the soil solution, 20-70% may exist in organically bound forms (Speir and Ross, 1978). Soil organic P compounds can be classified into three groups (Anderson, 1980): ( 1 ) inositol phosphate which composes up to 60% of soil organic P (Tate, 1984), ( 2 ) nucleic acids, and (3) phospholipids. Dalal (1977) and Tate (1984) reported that soils contained phosphoproteins, sugar phosphates, and glycerophosphates. A large portion of unidentified organic P occurs as insoluble comFlexes bound with clay minerals and organic matter (Tate, 1984). Forms of organic P resistant to both chemical and enzymatic hydrolysis are thought to result from the incorporation of organic compounds into humic materials during oxidative polymerization of polyphenols. lnositol phosphate, which can have mono-, di-, or triesters of phosphate groups, constitutes the most important fraction of organic P (Hawkes et a/.,
144
F. IYAMUKEMYE AND K.P. DICK
1984; Gil-Sotres et al., 1990). Some esters are quickly broken down in soil (Anderson, 1980); however, they can react and form stable complexes with Fe and A1 as organic P (Saxena, 1979) or be stabilized by colloids (clay, sesquioxides) through sorption reactions (Stewart and Tiessen, 1987). These may be the reasons why they tend to accumulate in soil compared to phospholipids (Cole et al., 1977). Another very important dynamic P pool is the biomass P which is 1 to 2% of the total soil P (Stevenson, 1986) and is directly correlated with biomass C (Brookes et a / ., 1984). Isotopic double-labeling techniques have shown that recently added organic residue P apparently is an important component of the microbial biomass P. A field experiment showed that 22-28% of the 33Papplied in medic plant residues was recovered in the microbial biomass (McLaughlin et al., 1988a). Furthermore, there appears to be rapid transformation of plant P to organic P fractions in soils. For example, McLaughlin er al. (1988b) reported that after 7 days 40% of the plant residue 33P was incorporated into organic P fractions of soil. The microbial biomass C represent 2-3% of the total organic C in soil (Sanyal and De Datta, 1991) but it is a key site for soil organic P mineralization (Brookes et a l . , 1984). An early incubation experiment of soil provided evidence that organic P is mineralized because organic P decreased similarly to increases in extractable inorganic P (Van Diest and Black, 1959). Other evidences of organic P mineralization are provided by observations that organic P decreases in cultivated soils (Sanyal and De Datta, 1991; Haas, 1961).
B. EARLYHISTORY Phosphorus fixation was first demonstrated in 1850 by Way (1850) with simple percolation experiments. Calcium phosphate dissolved in dilute sulfuric acid was passed through soils which resulted in leachate with no detectable levels of orthophosphate. This indicated that there was a rapid reaction of orthophosphate with soil constituents. Liebig (1858) and Sachs (1 865) demonstrated that polished plates of marble and ostheolite were etched by the roots of different plants. Later views held that plants excreted acids from their roots, which were proposed to be effective agents for solubilizing mineral forms of nutrients like P (Dyer, 1894; Quartaroli, 1905; Pfeiffer and Thurmann, 1896; Palladin, 1911; Maze, 1911). In particular, Shulov (1912) proposed that excretion of malic acid from plant roots solubilized orthophosphate in soils. Yet as late as 1931, Miller (1931) argued that there was no positive proof that carbonic acid was excreted by plant roots. Later Gerretsen ( 1948) provided evidence that microorganisms in the rhizosphere can solubilize insoluble mineral P forms.
ORGANIC AMENDMENTS AND PHOSPHORUS SORPTION
145
At the turn of the century, there was recognition that rock phosphate was best utilized when in intimate contact with decaying organic matter (Hopkins, 1910). It was thought that organic acids, carbonic acids, and nitrous acids resulting from decomposition were the active agents that were solubilizing P (Bauer, 1921). During the second decade of the 20th century there already were reports that organic amendments could affect P reactions in soils and plant availability. Jensen (1917) performed an experiment which included the use of alfalfa hay, sweet clover, barley hay, and stable manure added to soil. In addition, soils were extracted with: (a) soluble organic matter obtained from decomposing green manure and from stable manures; (b) soluble organic matter obtained from thoroughly decomposed green manure; (c) artificially prepared humus; and (d) osmosed organic solutions derived from the decomposition of organic matter. The conclusions were: 1. The amount of phosphoric acid dissolved by organic extracts from the soil exceeded the amount dissolved by water from 1.7 to 5.4 times regardless of the amendment; 2 . The solvent action of these organic extracts on the soil minerals appeared to be due both to the inorganic salts present in the organic solvents and to the organic compounds. 3. Mixing green plant manures or stable animal manure with soil (3% w/w) followed by incubation increased the solubility of phosphoric acid in the soils from 30 to 100%. 4. Artificial humus solution, free from calcium, magnesium, Fe, and phosphoric acid, increased the solubility of soil phosphoric acid.
Another approach to increasing the plant availability of inorganic P, recognized early in the 20th century, was to mix inorganic P with animal manure prior to incorporating in soil. Tottingham and Hoffmann (1913) found that mono- and tricalcium phosphate mixed with animal manure increased P uptake in barley grown in pots. Later, Midgley and Dunklee (1945) found that pellets formed from a mixture of phosphate and manure markedly increased P availability to plants. Increasing the pellet size increased P availability to the crop on highphosphate-fixing soils. Bear and Toth ( 1942) suggested that phosphate fixation can be greatly reduced by incorporating large amounts of organic matter into the soil. They explained that humic acid produced during organic residue decomposition reacts with soluble Al and Fe to form humates which are less soluble than orthophosphate. Like the hydroxyl ion, humic acids also function as a replacing agent for adsorbed phosphate and may be used as partial substitutes for liming material and organic matter (Bear and Toth, 1942). Copeland and Merkle (1941) found that soils receiving manure had lower phosphate adsorption quotients. They hypothesized that the biologically active
146
F. IYAMUREMYE AND R. P. DICK
manure either exerts a protective effect upon the soil mineral colloids or helps to release fixed phosphate or both. They observed that addition of animal manure had a large effect on P availability, but soils with high organic matter did not have a significant advantage over soils with lower organic matter in making P available to plants. From this, they suggested that it is the “biologically active” organic matter that is important in P availability in soils. Swenson el a!. (1949) demonstrated that several organic anions form stable complexes with Fe and Al. Humus and lignin were effective in replacing phosphate from the basic Fe phosphates, probably because of the formation of stable compounds or complexes between the active Fe and humus or lignin. Citric, oxalic, malic, tartaric, malonic, malic, and lactic acids were most effective in preventing P precipitation by Fe or A1 and citrate was the most active anion in preventing orthophosphate precipitation between pH 4 and 6 (Struthers and Sieling, 1950). Dalton et al. (1952) concluded that organic matter added to the soil as an amendment is effective in increasing the availability of soil orthophosphate. This was attributed to microbiological decomposition of organic amendments which releases metabolic products to form stable complexes with Fe and A1 in acid soils.
11. AEROBIC SOILS: ORGANIC ACIDS AND PHOSPHORUS SORPTION A. ORGANIC ACIDSIN SOILS When organic residues are added to soils, organic acids may be added directly to soils with the residue or be produced as by-products during decomposition of the residues by microbial activity. Organic acids commonly encountered in soil solution tend to be significantly lower in cultivated soils than in the same soil under native vegetation (Fox and Comerford, 1990; Hue et al., 1986). This is shown in Table I, in which a plantation site managed with monoculture pine is compared to a natural site which had mixed treelshrub vegetation. Hue et al. (1986) compared subsoils of cultivated and noncultivated soil on the same soil types and found that oxalic acid ranged from 4 to 22 in forested soil whereas in cultivated soil it ranged from < 1 to 3.4 p,M with similar trends among other organic acids measured (citric, malic, malonic, succinic, lactic, formic, and phthalic). Stevenson and Ardakani (1972) reported values as high as 3.7-5.0 and 0-1.0-4.0 mM of acetic and malic acid, respectively. Iyamuremye et al. ( I 99%) detected organic acids (oxalic, malic, maleic, malonic, succinic, formic and acetic acid) in the soil solution of five different types of soils. The concentration of these organic acids varied
ORGANIC AMENDMENTS AND PHOSPHORUS SORPTION
147
Table I Low-Molecular-WeightAliphatic Organic Acids Identified in the Soil Solution from a Pomona Series Soil Supporting Pine Stands in Alachua County, Florida (Adapted from Fox and Comerford, 1990) Stand
Horizon
Oxalic
Formic
Plantation I
A Bh A Bh A Bh
193 362 62 683 12s 313 293 1043
19 32 63 148 89 45 64 77 tr 19
Plantation 2 Plantation 3 Plantation 4 Natural I Natural 2
A
Bh A Bh A Bh
Natural 3
A
Natural 4
Bh A Bh
358 142 98 460 154 733 198 465
114
6 151
I37 9 5
Citric
Acetic
Malic
Succinic
tr
tr tr tr tr tr tr tr
tr tr tr tr
-
tr
-
tr
tr
tr
Note. All values are in pM. Trace amounts: peak identified at appropriate retention times but peak area not integrated. No peak identified at appropriate retention time.
with soil type and organic amendment treatment. Stevenson (1967) reported values of formic acid ranging from 2.5 to 4.4 mM in soil solution.
B. COMPLEXATION REACTIONSWITH METALS The formation of metal complexes by organic acids for such metals as A1 or Fe which readily react with orthophosphate would increase orthophosphate availability to plants. Hue et al. (1986) demonstrated this by showing that organic acids detoxified A1 effects in relation to plant growth, thus providing evidence of reduced A1 activity. They further found that organic acids varied in their affinity to form complexes and had stability constants of AI-organic acid complexes which decreased in the order citrate > tartrate = malate > salicylate. Carboxyl and hydroxyl functional groups are important in reactions between metals and organic acids (Huang and Violante, 1986). An example of this is shown by chelation of Fe with citrate with the following complexation.
148
F. NAMUREMYE AND R. P. DICK
In soils, a wide range of biochemical compounds have been identified in forming complexes with metals. These have included aliphatic acids, amino acids, phenolic acids, hyroxamate siderophores, 2-ketogluconic acid, and polymeric phenols (Stevenson, 1994). The ability of FA to form stable complexes with metal ions can be attributed to their high number of oxygen functional groups such as carboxylic, phenolic, alcoholic, and enolic OH, (Tan, 1986). Schnitzer (1969) distinguishes two types of reactions between humus and metal ions or hydrous oxides. A major reaction involves the simultaneous acidic COOH and phenolic groups, and a secondary reaction only involves the less acidic COOH groups. The complexation can be visualized as follows. AP+
+ RCOO-
* RCOOAP+
(7)
C. COMPETITION FOR SORPTION SITES The adsorption of organic compounds such as humic acid (HA), FA, and other organic acids on soil minerals such as A1 and Fe hydrous oxides or other clay minerals has been established and may lead to a competitive adsorption for sites of orthophosphate fixation but this effect varies with soil type. For example, Appelt et al. (1975) reported that HA or FA did not decrease P sorption on a volcanic soil at certain concentritions of HA or FA (Tables I1 and 111) because new HA-AI(0H)y complexes form and become new source of sites of P fixation. On other soil types organic compounds have shown a consistent reduction of adsorption sites. Hajra and Debnath (1985) showed that HA applied to soil decreased Fe bound P in soil treated with P fertilizer (Table IV), indicating that HA may release P from Fe compounds and/or prevent P precipitation. Evans and Russell (1959) showed that FA from a podzol was sorbed on lepidocrocite and goethite. Sibanda and Young ( 1986) studied competitive adsorption on goethite, gibbsite and tropical soils and reported that HA and FA competed strongly with orthophosphate for the sites of adsorption at low pH, and that HA decreased orthophosphate sorption by these soils (Figs. 3 and 4). Measurements showed that when orthophosphate adsorption was increased there was virtually no release of either HA, or FA into solution. Apparently the effectiveness does not rely exclusively on the occupation of adsorption sites by carboxyl groups but also by
ORGANIC AMENDMENTS AND PHOSPHORUS SORPTION Table I1 Effect of Humic Acid (HA) Adsorption on P Adsorption by a Puerto Octay Subsoil (Adapted from Appelt el al., 1975)" HA added (gP ~
HA adsorbed (gP
P added
(mmotP
~
~~
P adsorbed (mmoI)b
pH at equilibrium
6.0 12.0 17.5 23.0 28.7
5.7
~
0.0 0.0 0.0 0.0 0.0
6.0 12.0 18.0 24.0 30.0
3.8 3.8 3.8 3.8 3.8
6.0 12.0 18.0 24.0 30.0
~
-
I .4
5.8
5.9 6.0 6. I
5.8
6.0 12.0 17.2 23.0 28.5
1.4
I .4 1.4 1.4
5.9 6.0 6. I 6.1
( I Each number represents the mean value of two determinations with a fluctuation h Per 100 g of soil. <S%j.
Table 111 Fulvic Acid (FA) Adsorption (meq1100 g soil)" and Effect on P Adsorption (mmo1/100 g soil) by a Puerto Octay Subsoil, When Increasing Amounts of FA and 20 mmol of P Were Added at pH 4.0 or 5.5 (Adapted from Appelt eta/., 1975)h __
FA adsorbcd
P adsorbed ___ -
FAadded
pH40
pH55
pH40
pH55
0.0 5.0 10.0 15.0 20.0 25 .o
-
-
20.0 20.0 20.0 20.0 20.0 20.0
20.0 20.0 20.0 20.0 20.0 20.0
S .0
10.0 15.0 20.0 25 .0
5.0 10.0 10.0 13.0 16.0
Meq represents the amount of NaOH solution necessary t o titrate a dialyzed sample of FA from pH 35. to 8.0. Each number represents the mean value of two determinations with a fluctuation of <5%.
149
F. IYAMUREMYE AND R. P. DICK
150
Table IV Effect of Humic Acid on the Transformation of Added Phosphate (Adapted from Hajra and Debnath, 1985) Increase in P in various fractions (mg P kg-l)
NH.,CI sol. P Days of incubation CcJ HA HA-C ~~
AI-P
Fe-P
.
C
HA
HA-C
C
____
HA
HA-C
76 80 80 90
67 72 72 72
-9 -8 -8 -18
75 76 78 78
70 73 75 75
C
Ca-P
HA
HA-C
-
Red soil
30 60 120
180
tr tr tr tr
14
14
I1 8 9
11 8
tr tr tr tr
II 10
II
11 10
II
tr
6 4
6
9
88 87 84 75
90 88 87 84
2 1
3 9
8 8 9 10 9 10 8 12
0 1
-5 -3 -3 -3
6 4 4 4
10 7 8 8
4 3 4 4
-12
7 8 8 10
10 10
3 2
9 12
2
1
4
Laterite soil
30
60 120 180 30 60 120
180
tr tr tr
4 8
10 10
4 4
8
68 65 66 62 64 60 64 60 84 67 56 51
92 80 71 68
-3 -4 -4 -4
Alluvial soil 7 97 85 13 118 101 21 125 101 125 I08 17
-17 -20 -17
1
C, control; tr, trace; HA, humic acid-treated soil; HA-C, net extractable P for given P fraction (humic acid-treated soil minus untreated control).
an unfavorable electrostatic field generated around adsorbed HA molecules. Also, part of the energy of HA must be physical (van der Waals) in nature and would not be involved in competition with orthophosphate (Sibanda and Young, 1986). Organic acids such as citrate, malate, and oxalate can compete with P for sorbing sites (Struthers and Sieling, 1950; Hue, 1992; Violante and Gianfreda, 1993). Violante and Gianfreda (1993) have studied the competition between orthophosphate ions and oxalate and showed that on a montmorillonite (chloritelike) complex, more orthophosphate than oxalate was sorbed in a system containing a constant amount of orthophosphate even when the initial concentration of oxalate was higher than that of orthophosphate. Of the sites on the clay mineral that were available for adsorption by both anions, 51 to 79% of the sites were occupied by orthophosphates. Many sites were highly specific for orthophosphate, whereas others were common to both oxalate and orthophosphate, but these sites still had greater affinity for orthophosphate than oxalate. Yet other
ORGANIC AMENDMENTS
0.2 0.4
AND PHOSPHORUS SORPTION
0.6 0.8 1.0
1.2
1.51
1.4
PHOSPHATE IN SOLUTION (pg P cm-J)
PHOSPHATE IN SOLUTION (pg P mJ)
Fcgure 3 Phosphate adsorption isotherms on Salisbury (Harare) 5E2 series topsoils (0-20 cm), conducted at the measured pH of the soil in 0.1 M NaCl and at 25°C. (a) Salisbury (Harare) 5E2 series. and (b) Marandellas (Marondera) 7G2 series. Four levels of hurnic acid are compared: A, 0; B , 0.4%; C. 1.6%; D. 3.0% (adapted from Sibdnda and Young, 1986).
sites were specific for oxalate. Maximum reduction of adsorption of orthophosphate occurred when oxalate was added before orthophosphate addition and, conversely, the minimum occurred when orthophosphate was added first. Addition of phytic acid to soil strongly reduced P sorption whereas cinnamic and benzoic acid had no effect on P sorption (Evans, 1985). Cations bound to P by organic acids resulted in the release of P in the soil solution. In the case of oxalate, ligand exchange sites are complexed by the following reaction.
1
1 ,I
,AI:i
1
u-L
+
2OH-
F. IYAMUREMYE AND R. P. DICK
I52 f
6.0 r
P
(a)
0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6 1.8 2.0 2.2 2.4 PHOSPHATE IN SOLUTION (pg cm-7
U
Y
5.0
D
a:
0
$ a
4.0
3.0
W
48 0
2
2.0 1.0 d
0
0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6 1.8 2.0 2.2
PHOSPHATE IN SOLUTION (pg ~ r n - ~ )
Figure 4 Phosphate adsorption isotherms o n synthetic geothite, conducted in 0.1 M NaCl and at 25°C. (a) at pH 4 and (b) at pH 7 . Four levels of humic acid are compared: A , 0; B , 11.8%; C, 47.1%;D, 88.2% (adapted from Sibanda and Young, 1986).
The presence of certain organic anions can greatly reduce the amount of orthophosphate precipitated. Swenson et al. (1949) showed that humus and lignin had the ability to prevent orthophosphate fixation (Fig. 5). They concluded that humus and lignin were effective in replacing phosphate from the basic Fe phosphates, probably because of the formation of stable compounds or complexes between the active Fe and humus or lignin. Several organic anions including aromatic hydroxy acids and aliphatic hydroxy-organic acid were effective in preventing orthophosphate from combining chemically with A1 or Fe or in replacing the chemically combined orthophosphate. The pH has a major effect on competition between organic acid and P for adsorption sites (Violante et af., 1991; Hue, 1992; Lopez-Hernandez et a / . , 1986; Nagarajah et al., 1970). Struthers and Sieling (1950) showed that citrate decreased P sorption at any pH (Fig. 6 ) and concluded that the most effective acids were citric, oxalic, kartaric, malonic, malic, and lactic acids. Generally
ORGANIC AMENDMENTS W D PHOSPHORUS SORPTION
153
L
0.4
I
I
I
0.2
0.4
L
0.6
0.8
I
1.0
1.2
ORGANIC MATTER ADDED (9) Figure 5 Ability of humus and lignin to prevent phosphate fixation. Three milliequivalents of Fe,O,-Fe was reacted with I meq phosphate with or without organic amendments. Saloid-bound phosphate is amount of phosphate in solution when only Fe,O, and phosphate were in solution (adapted from Swenson er a / . . 1949).
3
4
5
PH
6
7
8
PH
Figure 6 Influence of pH on effectiveness of lactate. malate. citrate, and a-aminopropionate ions in preventing phosphate precipitation by iron ( a ) and by aluminum (b) (adapted from Struthers and Sieling. 1950).
154
F. NAMUREMYEC AND K. P. DICK
these acids were the most effective complexing agents at a more acid pH which is fortuitous because this coincides with soil pH values when Fe and A1 are most active in phosphate sorption, thus helping to prevent fixation.
D. DISSOLUTION OF PRECIPITATED PHOSPHATE AND FORMATION OF SOLIDPHOSPHATE PHASES Dissolution of precipitated phosphate by organic acids is another probable mechanism that makes P available when organic residues are added to soils. The solubilization of phosphate by organic acids involves the reactions of carboxyl and hydroxyl functional groups of organic acids with Al, Fe, and Ca cations to form complexes. The effects of organic acids on P availability starts at the early stage of soil formation. Organic acids are capable of solubilizing the phosphate minerals during the weathering process and therefore can release phosphate bound in these minerals. Bolton (1882) stressed the role of organic acids in the weathering process of rocks and subsequent release of nutrients. This ability of organic acids to solubilize P from minerals has been used to determine plant available P in different soils (Bolton, 1882; Morgan, 1941). The reaction of organic acids with soluble A1 results in the reduction of the activity of A1 and Fe and a corresponding increase of P activity in the soil solution as shown in the case of variscite which controls A1 solubility as follows (Lindsay ef al., 1959). Al(OH),H2P04 & A13+ + 20HpAl
+ H,PO,'-
+ 2,OH + ,H2P04 = pK,,
=
30.5
(9) (10)
Because the K s p remains constant, a decrease in A1 with no change in pH results in a release of phosphate. This means that additional variscite may solubilize. The reaction above shows that organic acids may prevent P precipitation or increase P concentration by solubilization of minerals. Organic acids also may react by ligand exchange with orthophosphate sorbed on the sites of Al, Fe, or Ca oxides and hydrous acids (Lopez-Hernandez et al., 1979). The kinetics of release seems to be related to the type of organic acid. Traina et al. (1986) reported a more rapid initial release of orthophosphate from an acid montmorillonitic soil in the presence of citrate compared with formic acid. The ligand that formed more stable complexes with A1 increased the rate of orthophosphate release. Recently, Fox et al. (1990) proposed two mechanisms for the role of organic acids in solubilization of P: (i) replacing P sorbed on metal hydroxides (Stumm, 1986), and (ii) dissolution of P sorbed at the metal-oxide surface (Stumm and Morgan, 1981; Martell et al., 1988).
ORGANIC AMENDMENTS AND PHOSPHORUS SORPTION
15s
The organic compounds may change physical properties of crystalline A1 or Fe. Schwertmann (1966) demonstrated that crystallization of A1 and Fe compounds was inhibited in the presence of HA and FA. Inoue and Huang (1983, 1984) demonstrated that complexing Al prevented the growth of A1 silicate crystals because of the disruption of the hydroxyl bridging, which is indispensable for the polymerization of A1 ions.
E. EFFECTSON SURFACE CHARGE Organic acids can have a significant effect on surface charge and precipitation reactions. Our review focuses on the literature for this topic as it relates to P sorption. Other reviews, such as Huang and Violante (1986), provide an in-depth discussion of the interaction of organics with soil minerals. Soil organic matter content affects the degree of surface charge. Indirect evidence of this relationship is provided by Moshi et al. (1974) who studied the distribution of charges and orthophosphate adsorption characteristics through entire soil profiles. They found that the negative charges were highest in the surface horizon of forest soils, and within a profile the negative charge decreased with depth, becoming almost constant below 30 cm in the cultivated profile and below 60 cm in the forest soil. Positive charge decreased with increasing organic matter which corresponded to a decrease in orthophosphate sorption. However, they suggested that the two phenomena are probably independent of each other, though this was not shown experimentally. Organic matter increases negative charges which may contribute to the decrease of P sorption by a repulsion of orthophosphate ions by a negatively charged surface (Moshi et a / . , 1974). Perrott (1978) found that treatment with extracts of humified clover caused the net surface charge to become more negative. Also, the positive charge on A1 silicates is reduced by organic treatments. The effect of organic treatment on charge characteristics of allophanic soil clay was similar to that of aluminosilicate clays. Perrott (1978) explained that the increase in negative charges in some soils was due to the selective adsorption of organic molecules containing a higher concentration of anionic groups. In the samples where this explanation holds, these authors suggested that the increase in negative charges was due to the removal of charges balancing Al hydroxy species, corresponding to the removal of A1 by organic matter from the structural models described by Cloos et al. ( 1968) and de Villiers (1971). Also, adsorption of organic matter at positively charged sites could expose negative charges by the conversion of the charge balancing A1 hydroxy polymers to insoluble organomineral complexes, similar to insoluble hydroxy-Al complexes (Hsu, 1968, 1979; and Huang and Violante, 1986). The organic matter may change surface
156
F. WAMUREMYE AND R. P. DICK
charges through the organic acids produced during mineralization or by their action on pH and exchangeable Al. Easterwood and Sartain (1990) reported a change in the nature of surface charges of soil samples previously coated with Fe oxides or with clover. Samples treated with clover had increased negative charges and resulted in more solution orthophosphate and less retention of P than in soil treated with Fe oxides.
F. PHYTOAVMLABILI-I-Y OF PHOSPHATE To confirm the increased P availability in the presence of organic acids, some bioassays have been conducted to determine the effects of organic acids on P phytoavailability. Hue (1991) showed that dry matter production of lettuce was increased when inorganic P was added with organic acids such as malic or protocatechuic acid (Fig. 7). Hue (1992) further concluded that the efficiency of P fertilizer would increase significantly when applied along with organic acids or particularly when acid-producing materials such as green manure or animal wastes were incorporated into soil. Moshi et al. (1974) reported that the amount of fertilizer P required to attain 0.2 mg liter' in the soil solution decreased from 90 to 22 rng kg-1 in an Oxisol when soil C increased from 3.8 to 6.5%.
111. AEROBIC SOILS: PLANT RESIDUES ANDANIMALMANURES
A. SOILPH Addition of organic residues can cause a significant increase in soil pH (Sharpley et al., 1984; Hoyt and Turner, 1975). Hue (1992) and Mnkeni and MacKenzie (1985) obtained higher pH values in soil solutions treated with organic residues than in those amended with CaCO,. Table V provides an example of pH changes due to organic amendments on some acidic soils from Oregon and Rwanda (lyamurernye er al., 1995a). The pH changes have been attributed to the high concentrations of basic cations in the organic matter used (Hue, 1992; Hoyt and Turner, 1975) and to the reduction of higher valence Mn oxides (Hoyt and 'Ibmer, 1975) or Fe oxides and hydrous oxides (Hue, 1992) in soils. The latter reaction leads to the self-liming effects observed in submerged sails caused by oxido-reduction reactions:
* Mn2+ + 20HFeO(0H) + e- + H,O * Fez+ + 30H- . MnO,
+ 2H+ + 2e-
( 1 1)
(12)
ORGANIC AMENDMENTS AND PHOSPHORUS SORPTION h
5 1.50
c.
h
g 1.50
3
'zpl
v
a
g 1.00 I= a
Ea
2 0.50
'.OO
2 0.50
z a: n
>-
a
f2 0.00
157
None Plot.
Mal.
None Prot.
Mal.
+ 100 mg P/kg
NO P
None Prot.
Mal.
None P r O l
None Prot.
Mel.
None Prot
Mal.
Mml.
+ 400 mg Plkg
NO P
None Prot.
+ 40 mg Pkg
NO P
0.00
MaI.
None Prot.
Mml.
+ 400 mg Plkg
NO P
Figure 7 Yield increases of lettuce in the presence of malic (Mal.)'or protocatechuic (Prot.) acid. Vertical bars are standard errors. Label " H " represents 2.0 mrnol/kg in the Oxisol or 5.0 mmolikg in the other soils; label "L" represents 0.5 mnlolikg in the Oxisol and 2.0 mmollkg in the other soils (adapted from Hue. 1991). Table V
The p H Values of Five Soils Treated with Organic or Inorganic Amendments after 28-Day Incubation (Adapted from lyamuremye et al., 1995a) Soil amendment -
~
~
~
Soil
Control
ManureU
Alfalfao
Wheat straw"
CaCO,*
CaSO,h
Jory (Ultisol) Mata (Ultisol) Tolo (Andisol) Kinigi(Andiso1) Kibeho(Ultiso1)
5.4 4.7 5.8 4.7 4. I
6.4 6.0 6. I 5.3 5.8
6.7 6.2 6.7 6.1 5.9
5.5 5.3 6.2 5.5
5.8 5.9 6.2
5.3 4.5
5.0
4.8 5.3
5.4 4.6 4. I
-
Amended to m1 at 5% ( w i w ) . Amended to soil at rate of three times the equivalents needed to neutralize exchangeable Al.
F. WAMUREMYE AND R. P. DICK
158
Theoretically, these reactions should not occur in aerobic soils. However, because of the heterogeneity of the bulk soils it may be possible that anaerobic microsites can occur in aggregates. Hue (1992) also suggested that pH may increase due to ligand exchange between organic acids and hydroxyl groups of A1 or Fe hydrous oxides as follows:
(gibbaiie)
(tartrate)
The increase in pH causes an increase in cation exchange capacity which results from formation of negative charges on colloidal fractions.
B. EXCHANGEABLE ALUMINUM AND IRON In acid soils, high levels of exchangeable A1 and Fe play a significant role in controlling orthophosphate concentration in the soil solution. Thus reduction of exchangeable A1 and Fe by organic soil amendments could have a significant effect on P sorption. Iyamuremye er al. (1995a) reported that organic amendments (animal manure and plant residue) reduced exchangeable A1 on high P fixing soils after a 28-day incubation (Table VI). From a greenhouse experiment (Fig. 8), it can be seen that the effects of organic material may be temporary, with effects lasting not longer than 3 months (Hoyt and Turner, 1975). However, this may be long enough to affect P availability to annual crops and substitute for lime.
Table VI
Exchangeable A1 in Soils Amended with Organic Residues (5% w/w) and Incubated for 28 Days (Adapted from Iyamurernye et al., 1995a) Soil amendment [cmol (+) kg-'1 Soil
Control
Manure
Alfalfa
Wheat straw
Jory (Ultisol) Mata (Ultisol) Tolo (Andisol) Kinigi (Andisol) Kibeho (Ultisol)
0.426 0.704 nd 0.515
nd" 0.009 nd 0.037 nd
nd 0.004 nd 0.033 nd
0.093 0.296 0.004 0.204 0.321
nd, not detectable
1.080
ORGANIC AMENDMENTS AND PHOSPHORUS SORPTION
159
I
0
1 2 3 4 5 INCUBATION PERIOD (months)
6
Wgure 8 Exchangeable A1 measured in soil over a 6-month period following addition of alfalfwneal, sugar, or peat moss at l .5 or 3.0 % of the soil weight: (a) 1.5% rate, (b) 3% rate. (The value at 0 months is for the soil without organic material added.) (Adapted from Hoyt and Turner, 1975.)
Several mechanisms may be operating in the reduction of exchangeable metals when organic residues are added to soils. These reductions may be due to precipitation of Al ions by OH ions released from the exchange of ligands between organic anions and terminal hydroxyls of Fe and Al oxides; and/or the complexation of Al by organic molecules (Hoyt and Turner, 1975; Hue, 1992). Hue (1992) calculated the lime potential of the organic material used in his experiment and estimated that the addition of 5 or 10 g poultry manure kg-I was equivalent to 3.39 and 6.74 cmol Ca(OH), kg-1. Organic acids such as oxalic, malic, malonic, and citric acid are believed to complex exchangeable Al (Hue et al., 1986) and were found to be in a greater concentration in forest soil than in agricultural soils. Iyamuremye et a/. (1995~)detected oxalic, malic, malonic, maleic, succinic, formic, and acetic acids in the soil solution samples treated with organic residues. Speciation modeling of manure-amended soil showed that citric acid had a major role in complexing Al and Fe and affecting P activity in the soil solution (Iyamuremye et al., 199%).
160
F. IYAMUREMYE AND R. P. DICK
C. PHOSPHORUS CONTENTOF ORGANIC RESIDUES AND PHOSPHORUS SORPTION Organic materials generally contain P. This P is susceptible to mineralization which would release orthophosphate into soil solution. In turn, this orthophosphate could affect P sorption capacity of soils. Singh and Jones (1976) observed that all organic residues decreased orthophosphate sorption, up to a 30-day incubation period. They also found that only organic residues with more than 0.3% P decreased orthophosphate sorption and increased desorption up to 150 days. High amounts of orthophosphate were sorbed after 75 days of incubation when total P content in organic residue was less than 0.3%. They concluded that P initially fixed began to reappear in the soil solution by the end of a 150-day incubation period and that use of sorption techniques for predicting P requirement should take into consideration the type and amount of organic matter added to the soil. The C:P ratio of organic residues added to soil apparently is important in determining the effect of organic residues on P sorption. Singh and Jones (1976) found that when the ratio C / P was > 130, the organic manures did not decrease P sorption. Figure 9 shows that some organic’materials low in total P (sawdust, wheat, and corn) increased P sorption in soils, whereas poultry manure, barley straw, and legume residue decreased P sorption in soils. This same study further showed that an incubation of 150 days with soil amended with poultry manure had significantly ,iigher desorption of P than incubation of sawdust-amended soils. However, Bumaya and Naylor (1988) found that plant residues with P contents greater than 0.1% applied to soils at a rate equivalent to or greater than 5% (w/w) increased the extractable P and decreased orthophosphate sorption in a high P sorption soil.
D. BIOLOGICAL TRANSJXXMATIONS OF PHOSPHORUS NVI) FATE OF PHOSPHORUS FROM ORGANIC AMENDMENTS Studying P nutrition in a pasture/cereal rotation system, McLaughlin and Alston (1986) found that of the total P applied to the soil, P from plant residues contributed approximately one-fourth of that supplied by fertilizer; of the total P in the wheat plant, P from residues supplied approximately one-fifth of that supplied by the fertilizer. In this study it was found that most of P added through plant residues accumulates in the microbial biomass pool. From the above discussions, it is apparent that some P added to soil via organic residues is converted to inorganic P. However, for mineralization to occur, soil organic amendments must contain at least 0.2% total P, otherwise net immobilization may occur (Tisdale et al., 1985; Dalal, 1977). The C:P ratio has
ORGANIC AMENDMENTS AND PHOSPHORUS SORPTION
161
600 150 DAYS SORPTION INCUBATION
500
~~~~
0 AL = Alfalfa A
BA=Barley
0 BE=Beans 400
% v
A H
0
+
300
n W
gm
CH=Check CO=Corn PM = Poultry manure SD=Sawdust WH=Wheat
200
a 100
1 0.01
1
t
1
1
0.1
I
I
1.o
P IN EQUlLlBRiUM SOLUTION (pg mL1)
Figure 9 Phosphorus sorption following incubation with organic residues for 150 days (adapted from Singh and Jones, 1976).
been used to predict P mineralization; it is generally accepted that C:P 300 leads to immobilization. One of the mechanisms advanced to explain the reduction of the adsorption capacity when soil is amended with organic residues is the complexation of orthophosphate sorption sites by orthophosphate added with or released from organic residues (Reddy et al., 1980; Iyamuremye et a/., 1995b). Guertal et al. (1991) reported that orthophosphate sorption was increased in surface soil samples where P had been extracted with resins. This showed that P from organic residues was occupying the sites of P adsorption. Both Chauhan et a/. (1979), with grass, and Sharpley et al. (1984), with feedlot wastes, found that labile and chemisorbed inorganic P increased when soils were amended with these materials. Iyamuremye el a/. (1995b) showed that manure and alfalfa increased labile and chemisorbed P in five high P-fixing soils which represents the P that has reacted with Al and Fe through precipitation and/or adsorption on surfaces of A1 and Fe compounds in acids soils. This indicated that soluble inorganic P added with organic residues and/or mineralized residue-P was reacting with the sites of P fixation. These two materials had
F. IYAMUREMYJZ AND R. P. DICK
162
Table VII Sorption Parameters Calculated from the Langmuir Equation in Five Soils Treated with Soil Organic (added at 5% w/w or Inorganic Amendments (Added at Three Times the Equivalents Needed to Neutralize Exchangeable Al) (Adapted from Iyamuremye el al., 1995a)
Soil amendments
Jory
Mata
Kibeho
Kinigi
Tolo
Manure Alfalfa Wheat straw CaCO, CaSO, Control
0,46f' 0.55e 0.88d 1.08~ 1.40b 1.58a
Affinity (k) (liters cmol-I) 1.03e 1.05d 3.18b 1.28de 1.02d 4.18b 1.61cd 1.47~ 2.98b 1.89bc 1.60~ 8.20a 2.37ab 2.82a 5.74ab 2.48a 2.33a 5.91ab
0.18b 0.18b 0.30ab 0.42a 0.32ab 0.37a
Manure Alfalfa Wheat straw CaCO, CaSO, Control
3.31b 3.40ab 3.37~ 3.35b 3.40ab 3.51a
Adsorption maxima ( b ) (cmol kg-I) 3.03b 3.43d 3.62a 3.24ab 3.59~ 3.62a 3.37a 3.65bc 3.62a 3.32a 3.66b 3.69a 3.37a 3.67b 3.74a 3.38a 3.97a 3.79a
2.89d 3.05~ 3.40a 2.78e 3.36b 3.31b
Manure Alfalfa Wheat Straw CaCO, CaSO, Control
0.76f 0.89e 1.20d 1.30~ 1.60b 1.77a
P sorbed at 0.2 1.20~ 1.40~ 1.70b 1.8Oab 2.00a 2.10a
mg liter-] (cmol kg-l) 1.4Od 2.40bc 0.30b 1.40d 2.60b 0.32b 1.8Oc 2.36~ 0.5% 1.85b 3.10~ 0.56a 2.40a 2.96a 0.58a 2.35a 3.00a 0.64a
I' Means followed by the same letter within a row are not significantly different (Tukey's test P < 0.05).
high P content in comparison with wheat straw which was less effective in increasing labile or chemisorbed P.These results are consistent with calculated P sorption parameters (Iyamuremye et al., 1995a), where manure and alfalfa had greater effects than wheat straw in decreasing adsorption maxima and affinity constants (Table VII). Mechanisms responsible for organic P mineralization are complex. Laboratory studies showed that growth of plant roots in soils causes reduction in organic P content, while organic P increased when phytase or phosphatase preparations were added to soil in the absence of plants (Jackman and Black, 1952; Thompson and Black, 1970); the absence of plants caused mineralization of native organic P (Thompson and Black, 1970). This suggests that the presence of plant roots
ORGANIC AMENDMENTS AND PHOSPHORUS SORPTION
163
and/or soil microflora results in mineralization of soil organic P, whereas the phosphatase preparation by itself serves as substrate for microflora and results in immobilization of P as organic P. Increases in P concentration in the rhizosphere soil solution are attributed to the hydrolytic cleavage of soil organic P by the action of extracellular phosphohydrolases (Tate, 1984). From these results, phosphatase activity was proposed as an index of P mineralization (Speir and Ross, 1978).
E. PHOSPHORUS SORPTION Early studies (Jensen, 1917) demonstrated that the solubility of Ca, Mg, Fe, and phosphoric acid in citrus soils of the Riverside district is measurably increased by the addition of green manure, stable manure, or their extracts. Other researchers (Sibanda and Young, 1986; Struthers and Sieling, 1950) have shown that addition of organic compounds to the soil prevented P adsorption. Organic compounds can block exposed hydroxyls on the surface of Fe and Al oxides (Appelt et a l . , 1975) and decrease P fixation capacity of these oxides (Dalton et a/., 1952). Gaur (1969) and Stevenson (1986) suggested that organic compounds form stable complexes with A1 and Fe which result in increased P solubility. Mnkeni and MacKenzie (1985) found a decrease in P sorption resulting from addition of plant residues or farm yard manure to upland topsoil and upland subsoil. However, they observed that the effects depended upon the nature of phosphate added (orthophosphate or polyphosphates). Reddy ef a / . (1980), studying soil utilized for animal waste disposal, reported that the soil that had received high rates of manure sorbed less P and desorbed more P (Fig. 10). This study showed that an increase in waste loading rates decreased the adsorption capacity of a soil and increased equilibrium P concentration (EPC) (which is the intercept value at zero P sorption). Estimates of EPC values were shown to be better related to plant growth than phosphate potential (Wild, 1950). On the P sorption isotherm in Fig. 10, P at zero sorption is much higher than 0.2 kg P ml- I , indicating that these soils have sufficient P available for plant growth [according to Fox and Kamprath (1970), 0.2 pg P ml- I provide 95% of maximum plant growih]. Desorption can also increase with organic amendments. Reddy et al. (1980) showed that the same soils amended with swine effluent increased P desorption with increasing P loading, but the amount desorbed depended upon the type of soil (Fig. 10). For example, at high loading rates, 41 and 23 mg P kg-I soil were desorbed by Norfolk and Cecil soils, respectively (Reddy et a / . , 1980). These researchers calculated P sorption parameters of the two soils as a function of loading rate and soil depth. The decrease in P sorption by high P-sorbing soils is indicated by the concomi-
F. IYAMUREMYE AND
I64
t
R. P. DICK
SWINE LAGOON EFFLUENT (a) 0-0
CONTROL
0-0
81 kg P ha’ yr‘
A-A
161
P IN SOLUTION (bg rnV)
br
m 3.
v
n W
m U
a w
0
0 3
U
P Pa
“1
SWINE LAGOON EFFLUENT (b) (FOUR EXTRACTIONS) NORFOLK (+Year Application) SERE/
30t // CECILSERIES
201
a cn 0
80
160
240
320
PHOSPHORUS APPLIED (kg ha-’ yr’)
Figure 10 (a) Phosphorus adsorption isotherm of the Norfolk soil (0 to 15 cm) as influenced by application of swine lagoon effluent for a period of 5 years; (b) The sum or P desorbed after four I-hr extractions with 0.01 M CaClz in Norfolk and Cecil soils (surface 15 crn) as influenced by loading rate of swine lagoon effluent (adapted from Reddy er al., 1980).
tant decrease of the adsorption maximum and the affinity constant (Reddy et al., 1980; lyamuremye et al., 1995a), and by the reduction of P required to maintain 0.2 mg liter-’ P in soil solution (Iyamuremye et al., 1995a). Evidence for this is shown in Table VII. In this case, organic amendments had a larger effect on the affinity constants and P sorbed at 0.2 mg liter-’ than on the adsorption maxima. Nonetheless, there was a consistent decrease in adsorption maxima due to organic residues in all five soils.
ORGANIC AMENDMENTS AND PHOSPHORUS SORPTION
165
The reduction of P sorption due to organic amendments in soils is undoubtedly the cumulative effect of several mechanisms. However, insights into the relative importance of these mechanisms was reported by lyamuremye et al. (199Sa) (Table VII), who compared organic amendments with inorganic amendments. Calcium carbonate decreased exchangeable Al and increased pH, and CaSO, also reduced exchangeable Al but did not affect soil pH. Both of these inorganic amendments had less of an effect on sorption constants than the P-rich organic residues (manure and alfalfa). Calcium carbonate and CaSO, did not increase Bray I P or total inorganic P, as did organic amendments. Iyamuremye et al., (199Sa) concluded that the production of organic acids or other organic complexing agents and effects on P fractions were more important than reduction in exchangeable metals in affecting P sorption when soils are amended with organic residues. Furthermore, wheat straw with a low P content also increased P in equilibrium solution and decreased sorption capacity in most of the soils studied, which suggested that other compounds or mineralization products in residues, such as organic acids, are involved in preventing P sorption.
F. ORGANIC AMENDMENTS ENRICHED WITH INORGANIC AND PHYTOAVAILABILITY OF PHOSPHORUS PHOSPHORUS There is extensive literature on organic amendments and plant nutrition but we are limiting our discussion to literature specifically related to effects of organic amendments on P sorption and P availability to plants. Pierre (1938) and Salter and Schollenberger (1938) discussed the general benefits of animal manure in P availability and crop productivity. When organic residues are added to soil, biological mineralization and production of organic P fractions are important in determining the availability to plants of P originating from organic residues. Results with wheat straw and wheat straw composts containing labeled P indicated that a relatively greater amount of P is utilized by plants from compost than from noncomposted material (Fuller and Nielsen, 1956). On the other hand, results with P-free extracted oat straw showed that the less decomposed material (oat straw plus KH,PO,) supplied more P to rye grass than did the compost (oat straw compost plus KH2P0,). The experiments also showed that indigenous soil P is made more available to plants as a result of decomposition of crop residues such as straw. Under field conditions Juang (1994) showed that simultaneous addition of rice compost and N-P-K fertilizers to soil resulted in extractable P levels in soils at harvest time for corn of 52 mg P kg-1 compared to a sole N-P-K fertilized treatment of 22 mg P kg- I . Injected dairy manure on a soil fertilizer equivalent basis showed that P availability to corn (Zea mays L.) varied widely (12-89% availability) as a function of location and time (Montavalli et a l . , 1989). This
F. IYAMUREMYE AND
166
R. P.DICK
---- MANURED PLOTS
-
UNMANURED PLOTS
0
150
450
600
POUNDS 16% SUPERPHOSPHATE
Figure 11 The relationship between total yield in bushels of corn, wheat, and oats (grain) and the pounds of 16% superphosphate applied per acre after 57 years of soil amendments and cropping. The upper curve represents plots receiving 6 tons of manure with added phosphate, the lower curve represents plots which received 125 pounds of NaNO, and 100 pounds of KCI but no manure. Plots receive varying amounts of 16%superphosphate as indicated. The average adsorption quotient for the manured plots is 0.9246 and for the unmanured plots 0.9510 (adapted from Copeland and Merkle, 1941).
high variability is likely related to the importance of biological activity in releasing organic P as a function of soil type and climatic factors. In an alkaline soil Abbott and Tucker (1973) showed that the main benefit from animal manure to crops was in P availability (as shown by nutrient content in cotton or barley tissue). They reported residual effects of P availability for up to 4 years when 22 tons ha-' of animal manure was applied to soils. Early work by Copeland and Merkel (194 l ) , as illustrated in Fig. 11, showed that inorganic P could be used more efficiently on soil that had received longterm application of animal manure. A logical extension of this work would suggest that mixing manure and P fertilizer may have the same effect on the yield. Indeed, Midgley and Dunklee (1945) reported that if phosphate fertilizers were added to soils after premixing with farmyard manure, the availability of P to plants was markedly better than if these materials were added separately. Later, Widdowson and Penny (1968) showed that a combination of farmyard manure and inorganic P produced the highest yields of wheat and barley. Giardini
ORGANIC AMENDMENTS AND PHOSPHORUS SORPTION
167
et a / . (1992) showed that mixing poultry manure and fertilizer produced greater crop yield than did manure or fertilizer alone for some crops. One early explanation was that biologically active manure either exerts a protective effect upon the soil mineral colloids or helps to release phosphate, or both (Copeland and Merkle, 1941). Much later, Sharif et al. (1974a,b) showed that premixing inorganic P with farmyard manure before addition to soil markedly increased the uptake of inorganic P by plants and increased crop yield on P-deficient soil. These authors hypothesized that organic matter may increase the availability of inorganic P by suppressing the conversion of inorganic P to Less soluble compounds such as Ca phosphates in calcareous soil. Over the years, a number of studies have shown that mixing farmyard manures and inorganic P significantly increased extractable P in soils (McAuliffe et a / . , 1949; Formoli and Prasad, 1979; Meek et al., 1979; Geiger et al., 1992). Plant availability of P from rock phosphate has been enhanced when mixed with organic material such as compost. Rock phosphate is likely solubilized by biological activity and production of acids during composting. A number of studies in India showed that mixing rock phosphate with animal manures or plant residues followed by composting increased citrate-acid soluble P and/or P availability to plants (Mishra et a l . , 1982; Mathur el al., 1980; Rastogi et al., 1976). In an alkaline soil, Singh and Amberger (1995) showed that compost enriched with rock phosphate resulted in greater P uptake by rye grass in the greenhouse and lower fixation with native Ca than soil amended with soluble inorganic P.
IV. WATERLOGGED SOILS A. ORGANIC AMENDMENTS AND EH The behavior of P in flooded soil is quite different from that of upland soils and has been extensively reviewed by Ponnamperuma (1972) and Yu (1985). Recently Sanyal and De Datta (1991) summarized the chemistry of P transformation in these soils. If the soil oxygen content is low, a series of soil physicochemical properties such as oxidation-reduction potential, pH, and the forms of some chemical elements capable of participating in oxidation-reduction reactions ( N , S, Mn, Fe, etc.) will be changed (Pan, 1985). Consequently, reduction is the main chemical reaction driving other biological reactions in saturated soils (Ponnamperuma, 1975). As the oxygen content decreases, the oxidation-reduction potential drops at the same time. The oxidation-reduction (Eh) system can be explained as a chemical reaction in which electrons are transferred from a donor to an acceptor. Eh is the most important index for characterizing the degree of oxidation or
168
I;. IYAMUKEMYF. AND K. 1'. DICK
reduction of a soil and reflects the equilibrium position among various redox systems (oxygen, iron, manganese, nitrogen, sulfur, carbon, etc.) when they exist in steady equilibrium (Liu, 1985). The free energy of oxidation and reduction reactions written as a function of reducing product (red) and oxidizing product (ox) can be written as follows,
AG
=
AC"
(red) (ox) '
+ RT In
~
where AG" is the standard free energy when (red) = (ox). In volts (V), Eq. (14) becomes Eh
=
E,,
(ox) + RT - In nF (red) ~
'
Other equations have been proposed (Sillen, 1964; SillCn, 1967a,b; Stumm and Morgan, 1970). and the most common one is I pE= - I n k , N where k is the equilibrium constant. The redox potential values needed for Fe phosphate compounds to be reduced is intermediate among reducible elements and may be reached easily in flooded soil (Ponnamperuma, 1972). During anaerobic respiration, organic matter is oxidized and soil components are reduced; the organic matter provides a source of electrons for biological reactions (Ponnamperuma, 1972). Therefore, addition of organic matter as a source of oxidizable material for microbial activity accelerates reduction (Willett and Higgins, 1978), and increases the rate of reduction in soil (Ding and Liu, 1985).
The presence of native or added organic matter hastens the reduction to the initial minimum equilibrium Eh value (Ponnamperuma, 1955; Yamane and Sato, 1961) which is due to a rapid release of C providing energy for reduction reactions. Figure 12 shows that addition of organic residues can cause a more rapid reduction in Eh but this is a function of the native organic matter content of the soil. The initial abrupt fall in oxidation-reduction potential results from the rapid consumption of oxygen by aerobic microbes. Nagarajah et ul. (1968) reported that the changes in Eh were more pronounced when organic substrates are added to soils low in organic matter.
B. ORGANIC AMENDMENTSAND PH When a soil is submerged, its pH decreases during the first few days (Motomura, 1962; Ponnamperuma, 1965), reaches a minimum, and then increases
OK<;ANICAMENDMENI'S AND PHOSPHORUS SORP'l'ION 300 200 100 v
c
w =!
o
0 v>
-100
169
THAKURGAON SOIL
:L
A-A
CONTROL
0-0
O
RICE STRAW
U PEAVINE
----I3
-200
----o-----o
----~-----o
-300
300 -
-
A
.
A
A-9
JOYDEBPUR SOIL
Rgure I2 Etkct of iidded organic residues on thr Eh of (a) Thakurgaon and (h) Jory soils incuharcd under I1c)twlcd conditions (adapted from Islam. 1993).
asymptotically to a final value a few weeks later. The largest increase in pH of acid soils occurs during the first 2 weeks of flooding followed by a gradual increase up t o 12 weeks in the absence o f organic amendments (Islam, 1993). Conversely, on alkaline soils pH decreases with flooding (Islam, 1993; IRRI, 1966; Ponnamperunia, 1965; Cang ct N I . , 1985). There are potentially several mechanisms involved in the pH change associated with flooded soils. Ding and Liu ( 1985) proposed that H ions participate in various redox systems. Another explanation is that the development of reduction conditions causes rapid decomposition of organic matter which releases OH-, consuming HI (Cang, 1985). The same authors, however, indicated that the intense decomposition o f organic matter, especially if the residue is easily de+
170
F. IYNMUKEMYE: AND R. P. DICK
composable, yields products such as organic acids which can temporarily retard the rise of soil pH, and even lower the pH. The increase of pH in acid soils results from production of OH-, whereas a decrease in pH of alkaline soils is due to accumulation of CO, which produces carbonic acid (Ponnamperuma, 1972). Soils high in organic matter and in reducible Fe attain a pH of about 6.5 within a few weeks of submergence, whereas acid soils low in organic C or in active Fe increase in pH more slowly and reach a lower equilibrium pH (IRRI, 1966). Singh ef ul. (198 I ) conducted an experiment with Azolla/blue-green algae composts or manure as amendments and reported that the flooding of soil increased soil pH gradually and reached a maximum at 50 days of incubation. The average pH in soil amended with fresh organic matter ranged from 6.61 to 6.69, whereas soil amended with compost or animal manure reached pH values of 6.59 and 6.43, respectively. Under anaerobic conditions, Fez+ forms. This dissolution of Fe is increased in the presence of organic matter. The reduction of Fe has important chemical consequences affecting pH changes and/or the solubility of P (Ponnamperuma, 1972). Ponnamperuma (1972) reported that acid soils high in organic matter had Fe concentrations as high as 600 ppm within 1 to 3 weeks of submergence. In turn, water-soluble F C +diffuses ~ to the oxygenated soil-water interface and can coprecipitate with P. Biological activity can affect Fe reduction directly or indirectly (Motomura, 1963). The former occurs when various organic C compounds produced in the course of anaerobic fermentation of organic matter dissolve Fe3+ in soil to become water-soluble Fe complexes or chelates. The latter occurs when reducing substances are produced by the biological reduction of Fe from ferric to ferrous forms.
C. FLOODED SOILSAND PHOSPHORUS SOLUBILITY Einsele (1936) and Mortimer (1941) were perhaps the first to report the increase in solubility of P under reducing conditions in lake sediments. Further evidence for this has been shown in lowland rice, which generally requires considerably less P fertilizer than upland crops grown on the same soil type (Ponnamperuma, 1965). These increases have bcen attributed to the solubilization of P following the reduction of such P compounds as Fe and A l phosphates (Sah et u l . , 1989a,b). Early evidence for this was the work of Fujiwara (1950) who found that lowland rice responded much better to Fe and Al phosphate than did barley or upland rice. Iron (111)-bound phosphate apparently is an important mineral involved in increasing soluble P in submerged soils (Ponnamperuma, 1955; Patrick and Wyatt, 1964; Savant and Ellis, 1964) by the following mechanisms:
0KC;ANIC AMENDMENTS AND PHOSPHOKUS SOKITION
I71
I. probable hydrolysis of Fe (111) and Al phosphates;
2. release of P held by anion exchange on clay and hydrous oxides of Fe; 3. reduction of Fe (111) to Fe (11) with liberation of sorbed and chemically bound P. The first two reactions are due to the pH increase brought about by reduction in soils. In alkaline soils, the increase in solubility of P is a consequence of the decrease in pH of these soils. For the solubility of hydroxyapatite, this increases as pH decreases (Stumm and Morgan. 1970). The phosphate released by these mechanisms and from the decomposition of organic matter may be resorbed by clay mineral and hydrous oxides of Fe in the anaerobic zone (Gasser and Bloomfield, 1955; Bloomfield, 1967) or it may diffuse to oxidized zones and be reprecipitated (Sah er d.,1989a). Sah and Mikkelsen (1989) and Sah et al. ( 1989a,b) reported that the forms of Fe in flooded-drained soils are influenced by the organic matter content of soil, and under some conditions amorphous Fe is formed at the expense of other forms so that the availability and sorptivity of P in flooded soils are correlated with Fe transformations (Table VIII). Wetting/drying cycles generally decrease the solubility of both native and fertilizer P (Brandon and Mikkelsen, 1979; Sah and Mikkelsen, 1989; Sah et al., 1989a,b). The common explanation is that the biological reduction of Fe during the flooding phase followed by reoxidation during the drying phase results in enhanced reactivity of the sesquioxide fractions of soil, leading to increased orthophosphate sorption. Islam (1993) found increases in P solubility with rice straw addition and a reduction with pea vine material (Fig. 13); the degree of soluble P measured varied with soil type.
D. EFFECTS OF ORGANIC AMENDMENTS ON PHOSPHORUS SORPTION Sanyal and De Datta (1991) have summarized the role of organic matter on P reactions in flooded soils. As in aerobic soils, organic acids may be released in soil solution following anaerobic decomposition of organic matter (Tsutsuki and Ponnamperuma, 1987). According to Willett (1986) these organic acids, when produced in calcareous soils, increase the solubility of Ca-phosphate compounds by complexing Ca ions. In acid soils, the complexation of soil Fe and soil A1 by the decomposition products of organic matter in Hooded soils may be responsible lor P release from insoluble compounds (Mandal and Mandal, 1973; Debnath rt d.,1974; Mandal. 1979). The role of mineralization of organic P in flooded soils is not well established. Tate (1984) suggested that it is relatively slow, and others have indicated that it is a minor source of P in flooded soils (Patrick and Mahapatra, 1968; Uwasawa et
172
F. IYAMUREMYE AND R. P. DICK Table VIII
Correlation of Changes in Fe Fractians with P Sorption as Affected by Organic Matter (OM) and Temperature Trehtments (Adapted from Sah et al., 1989b) Soil"
OM (g kg-I)
Temperature ("C)
AFeA
AFeB
AFeD
r with P sorption to maintain 50
Gridley (Typic Argixeroll)
0 10
10
Myers (Entic Chromoxerent)
0
10 10
Stockton (unclassified)
0 10 10
Yokohl (Typic Durixeralf)
I Ob 10
23
final P concentration (P,) 0.96 0.87 -0.99
23 35 All treatments? 23
0.98 0.95 0.93 0.89
I .oo 23 35 0.91 All treatments: 0.92
23 23 35 All treatments: 23 35 All treatments:
0.61 0.53 0.58 0.87
-0.94 -0.50 -0.63 -0.21
0.94
-0.81 -0.91
0.85 0.84
-0.94
r with F', at I .2 m M initial P concentration (P,) -0.67 0.88 -0.51
-0.70 -0.50 -0.64 -0.93
0.78 0.81 0.82 -0.59
-0.99 -0.94
-0.86 -0.70
+0.23 -0.24 -0.16 0.72 0.88 0.81
Nore. The variation in these two variables was brought about by different Rooding periods (FP). " Combined data of the combinations of OM and temperature listed for each soil. Correlation coefficients for Yokohl at 0 added OM are not shown due to lack of response of P,to
FP.
al. , 1988). Other researchers reported that organic amendments increased mineralization rates of organic P in flooded soils (Islam and Mandal, 1977; Goswami and Banerjee, 1978). Uwasawa et al. (1988) and Willett (1989) reported that P release in flooded soils originated more from P mineralization than from the solubilization of Fe components by reduction processes. Early on Paul and DeLong (1949) showed that flooding followed by drainage resulted in greater P fixation than the same soil prior to flooding. Furthermore they showed that easily decompasable organic matter increased P fixation and that the increased capacity for P fixation originated during the flooding stage, which they attributed to an increase in reactivity of the sesquioxide fraction.
ORGANIC AMENDMENTS AND PHOSPHORUS SORPTION
173
THAKURGAON SOIL
100
-
75 -
zn
CONTROL RICE STRAW PEAVINE
A-A
-
U-U
P
""I 0
2
0-0
4
6
8
10
12
TIME (weeks) Figure 13 Cumulative amount of PO,-P released in a Thakurgaon soil amended with organic residues under flooded conditions (adapted from Islam, 1993).
The Fe forms in flooded soils have been related to P sorptivity. Willett and Higgins ( 1978) investigated this relationship before, during, and after flooding with addition of rice straw. Their findings showed that upon flooding there were large increases in acetate- and oxalate-extractable Fe (Fig. 14) and P sorptivity. Furthermore, they showed a close relationship between organic acid concentrations and P sorption, as shown in Fig. 15. In the same study, oxidation of previously flooded soils resulted in decreased levels of P sorptivity and oxalate Fe, but at higher levels than before reduction. These authors concluded that P sorption processes by ferrous hydrous oxides dominate during the flooded (reduced) phase, but with reoxidation poorly crystalline ferric hydrous oxides dominate P sorption reactions. Previously, Khalid et al. (1977) concluded that P sorption could be related to oxalate-soluble Fe in flooded soils and proposed that ferrous hydroxides formed on reduction governed P sorption. Roy and De Datta (1985) came to a similar conclusion in wetland rice soils. The effect of organic matter was explained by a greater reduction of Fe (111) and the appearance of Fe (11) which caused a corresponding increase in P sorptivity. Khalid et al. (1977) also concluded that the greater fixation of added P in flooded soils and sediments may be described in
F. IYAMUREMYE AND R. P. DICK
174
1.2
8
v
z
E1 s0 a
1.0
0.8 0.6
[r
k
0.4
P
3 0
o.2
0
20
40
60
80
100
160
140
120
180
200
220
DAYS
Figure 14 The effects of flooding treatments on oxalate extractable iron in Willbriggie clay loam. (0)Control, not flooded; (A) flooded; (0) flooded with organic matter; ( 0 )flooded with rice; (A)flooded with organic matter and rice; 4 indicates day of drainage. Bars indicate the least significant differences ( p < 0.05) between the control and the mean of all the Hooded treatments (adapted from Willett and Higgins, 1978).
,-
h
b
' 600 0 °
1
8
0
A
500
8
A
Q
c8
300
200
0
200
400
600
800
1000
1200
1400
1600
ACETATE EXTRACTABLE IRON (pg gl)
Figure 15 The relationship between P sorbed and acetate extractable iron in reduced soils (Days 21-146) (adapted from Willet and Higgins, 1978).
ORGANIC AMENDMENTS AND PHOSPHORUS SORPTION I.
II. P, A
0 \>
b
A
lOaM 5ObM
lot1
.&(
3 8 1 4 79 1w~M 6 o t 6 75 ZW&M 83t.6 70
nlOl 383 396
411
A
A 0
a
b
17+04
64
60109
81 78
iar
73
t573
ag+og 11 8+12
111.
ulC
A
b
1.9t.2 .87 0 6.6t.4 .83 9.8t.4 79 1 2 5 ~ 4 .75
1104 12-30
175
cxlol
A
216.4
.)
247.8 268.7 268.5
..... ...............a ........... a
435
10
.............:
6
.............
0
20
40
60
00 100 0
... ..........
.......b . . . ........ .
...... ...........
a
20
40
60
00 100 0
20
40
60
80 100
DAYS OF FLOODING Figure 16 Effects of flooding period, organic matter (OM), and temperature treatments during flooding on P sorption to maintain an equilibrium (final) P concentration ( P , ) in flooded-drained Gridly soil (pH 5.2). Incubated under flooded Conditions: (I) Without added OM at 23°C; (11) With 10 g OM kg- soil at 23'C; (111). With 10 g OM kg- soil at 35°C. Regression model: Y = A( 1 - b exp-c"), where A = maximum expected P sorption. b and c are coefficients, x = days of flooding, and Y = expected P sorption (adapted from Sah and Mikkelsen, 1989).
terms of sorption involving a mixture of very reactive amorphous Fe(OH),Fe(OHI3 rather than precipitation of ferrous phosphate compounds such as strengite Fe,(P0,)*8H20. According to Sah et al. ( 1989a), organic matter and temperature are important factors in determining the impact of flooding-draining cycles on P sorption. They found that organic amendments and elevated temperature during flooding further increased P sorption and shortened the flooding period required to reach P sorption maxima. The soils in which P sorption did not increase without organic amendments were either low in organic matter or reducible Fe. However, the largest increase in P sorption as a result of organic amendments occurred in soil with low organic matter and high levels of reducible Fe when amended with organic residue (Fig. 16). Desorption of P is affected by organic matter addition in soils under floodingaerobic cycles. Sah er a / . (1989b) outlined several mechanisms that may be involved in these reactions and concluded that: 1 . Induced P deficiency under flooded-aerobic conditions may be expected in soils high in reducible Fe and containing a source of readily decomposable organic C . 2. Flooded-aerobic conditions increased P sorption and decreased P desorp-
E IYAMUREMYE AND R. P. DICK
176
(4
500
500
= .8 ,-
b CI)
c '
.s
400
.-cnb
o)
300
s.
I
400
300
v
ow200
m
a
5)
[I
2
100
n
100
n 0
1
2
3
4
5
6
P IN EQUILIBRIUM SOLUTION (pg P ml I )
0
1
2
3
4
5
6
P IN EQUILIBRIUM SOLUTION (pg P mil)
figure 17 P sorption of sandy loam soil amended with different green manures and incubated under flooded conditions for 20 days (a) or 40 days (b). (a)control; (A)sunhemp; ( X ) dhaincha; and (0) cowpea (adapted from Hundal ef a / . , 1988).
tion in soils where Fe transformations were not limited by either low organic C or reducible Fe content. 3. Organic matter treatments prior to flooding and elevated temperature decreased P desorption; there were much larger effects in soils with low organic matter but high reducible Fe content. Green manure incorporation significantly reduced P sorption capacity of anoxic soil (Hundal et al., 1988) as shown in Fig. 17. These amendments also reduced Langmuir parameters in soil incubated under flooding conditions (Table IX). The explanation given was that plant residues release P during the mineralization or that there was accumulation of intermediate organic acids that complexed metal cations. From the above discussion it can be seen that the relationship between organic residue additions in flooded soils have given conflicting results. Considerably more research is needed in this area, particularly with strict control of experimental conditions (e.g., redox potentials when measuring P sorption).
V. RESEARCH NEEDS In consideration of the current move toward sustainability and environmental stewardship, P is a key nutrient in maintaining long-term productivity of agricultural systems. Phosphorus as a fertilizer has finite world reserves and among
177
ORGANIC AMENDMENTS AND PHOSPHORUS SORPTION Table IX
Langmuir Parameters ( k and b ) for Sorption hy Soil Incubated with Three Green Manures for 20 and 40 Days under Flooded Conditions (Adapted from Hundal et al., 1988) 20-day incubation ~~
40-day incubation
~-
Green manure
k"
bh
k
b
Cowpea Dhaincha Sunhemp Unmanured control
0.037 0.044 0.058 0.120
0.375 0.400 0.462 0.600
0.036 0.043 0.053 0.070
0.333 0.353 0.387 0.522
Bonding energy (ml mg-' P). P adsorption maxima (mg P g - ' soil).
the nutrients has relatively low rates of recovery by plants when added to soils. After extensive review of the literature, we conclude that organic residue management can have a significant effect on P transformations and availability to plants under aerobic conditions. However, under waterlogging conditions and particularly under flooding-aerobic cycles, limited research has provided mixed results on the role of organic amendments in affecting P sorption. Although there is considerable evidence that organic amendments can decrease P sorption in acid soils, we were struck by the dearth of field-based research to determine whether the results from more basic studies can be applied under field conditions. Specific areas of research needed on the role of organic soil amendments and P sorptiodplant availability in soils are: 1. Investigations into partitioning the effect of the P contributions (inorganic or organic) vs the effect of organic acids on P sorption when organic residues are added to soils. 2. Determination of the transformations and fate of inorganic P when it is premixed with organic amendments prior to soil incorporation. 3. Studies on effects of organic matter addition on P sorption in in situ-flooded soils. Most studies have been done on flooded soils that were subsequently airdried prior to characterization of P sorption. 4. Assess potential for field applications for utilization of organic amendments to decrease P sorption and increase P fertilizer efficiency relative to plant nutrition. 5 . Establish conceptual and working models that have practical applications
178
F. NAMUREMYE AND R. P. DICK
and predictive capabilities for relating residue quality and soil environments to potential impacts on P sorption in soils.
REFERENCES Abbott, J. L., and Tucker, T. C. (1973). Persistence of manure. phosphorus availability in calcareous soil. Soil Sci. Soc. Am. Proc. 37,60-63. Anderson, C.(1980). Assessing organic phosphorus in soils. In “The Role of Phosphorus in Agriculture”(F. E. Khasawneh, E. C. Sample, and E. J. Kamprath, Eds.), pp. 41 1-431. Soil Sci. SOC. Am., Madison, WI. Appelt, H., Coleman, N. T.. and Pratt, P. F. (1975). Interactions between organic compounds, minerals, and ions in volcanic-ash-derived soils: 11. Effects of organic compounds on the adsorption of phosplrate. Soil Sci. Sor. Am. froc. 39, 628-630. Bauer, F. C. (1921). The relation of organic matter and the feeding power of plants to the utilization of rock phosphate. Soil Sci. 12, 21-41. Bear, F. E . , and Toth, S. J. (1942). Phosphate fixation in soil and its practical control. Indus. Eng. Chem. 34, 49-52. Berkheiser, V. E., Street, J. J . , Rao, P. S. C., and Yuan, T. L. (1980). Partitioning of inorganic orthophosphate in soil-water systems. In “CRC Critical Reviews in Environmental Control” (C. P. Straub, Ed.), Vol. 10, pp. 179-21 1. CRC Press, Boca Raton, FL. Bloomfield, C. (1967). Effect of some phosphate fertilizers on the oxidation of elemental sulfur in soil. Soil Sci. 103, 219-223. Bohn, H. L., McNeal, B. L., and O’Connor, G. A. (1979). “Soil Chemistry.” Wiley, New York. Bolton, H. C. (1882). Application of organic acids to the examination of minerals. froc. Am. Assoc. Adv. Sci. 31, 271-275. Brandon, D. M.,and Mikkelsen, D. S. (1979). Phosphorus transformations in alternately flooded California soils: I. Cause of plant phosphorus deficiency in rice rotation crops and correctional methods. Soil Sci. Soc. Am. J . 43,989-999. Brookes, P. C., Powlson, D. S., and Jenkinson, D. S. (1984). Phosphorus in the soil microbial biomass. Soil Biol. Biochem. 16, 169-175. Bumaya, A. H., and Naylor, D. V. (1988). Phosphorus sorption and extractability in Andic soil incubated with plant residues of variable P content. Plan1 Soil 112, 77-81. Cang, D. Q., Jing, H. W., and Zdng, X.-N. (1985). Acidity. In “Physical Chemistry of Paddy Soils” (Yu, T.-R., Ed.), pp. 131- 154. Springer-Verlag (Science Press), Beijing. Chauhan, B. S..Stewart, J. W.B., and Paul, E. A. (1979). Effect of carbon additions on soil labile inorganic, organic and microbially held phosphate. Cun. J. Soil Sci. 59, 387-396. Cloos, P., Herbillon, A., and Echeverria, J. (1968). Allophane-like synthetic silico-aluminas. Phosphate adsorption and availability. 9rh Inr. Congress Soil Sci. 2, 733-743. Cole, C. V., Innis, G . S., and Stewart, 3. W.B. (1977). Simulation of phosphorus cycling in semiarid grasslands. Ecology 58, 1-15. Copeland, 0. L.. and Merkle, F. G. (1941). The influence ofcertain soil treatments upon the fixation and availability of applied phosphates. Soil Sci. Sor. Am. Proc. 6, 321-327. Dalal, R. C. (1977). Soil organic phosphorus. Adv. Agron. J. 29, 83-117. Dalton, I. D., Russell, G . C.,and Sieling, D. H.(1952). Effect of organic matter on phosphate availability. Soil Sci. 73, 173-181. Debnath, N. C., Banerjee, S . K., and Mendal, S. K.(1974). Influence of humic and fulvic acids on the availability and inorganic transformation of phosphorus in soil. J . Indian Chem. Sac. 51, 81 1-813.
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ADVANCES IN DROUGHT TOLERANCE INPLANTS John S. Boyer College of Agriculture and College of Marine Studies, University of Delaware, Lewes, Delaware 19958
I. Introduction 11. Water Use Efficiency A. Physiology of Water Use Efficiency B. Importance of Harvest Index C. Measuring Water Use Efficiency D. Molecular Biology and Water Use Efficiency 111. Drought Tolerance A. Approaches to Improving Drought Tolerance B. Examples of Improved Drought Tolerance N . Water Deficits and Reproduction A. Mechanisms of Losses V. Desiccation Tolerance A. Molecular Features of Desiccation Tolerance VI. Conclusions References
I. INTRODUCTION The traditional solution to agricultural water shortage is irrigation. A steady supply of water makes agriculture possible in many otherwise nonproductive areas and the water often can bring reliable, high productivity. Because production becomes more predictable, investment can be made in other favorable cropping practices that result in further improvements in productivity. As a consequence, a major share of the global production comes from irrigated farms. Nevertheless, large supplies of water are necessary because most of the water is evaporated by the crop. The water is consumed and not returned for other uses, and in the United States more water is consumed by irrigation than by all other uses combined ( U . S. Department of the Interior, 1977). New supplies in the 187 A,ll,”nc~e,6, rllpn,tov,y iblrlmr $6 Copyrrghr 4 IOV6 by .Academic Prcss, Inc. .All rightq of rcpmducrion in any form reserved.
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United States have become scarce as municipalities and environmental needs compete for the same water. Salt-laden water usually is not an alternative because evaporation removes the water and not the salt, degrading the soil. As a consequence, new irrigation is becoming less possible than in the past and there is increasing interest in improving the efficiency of water use in irrigation and determining whether plants can yield well under water deficient conditions (Boyer, 1982). A number of methods exist for improving the efficiency of water use and have been summarized by Taylor et al. (1983) and by Stewart and Nielsen (1990). The methods can be classified in three broad categories: (1) increasing the efficiency of water delivery and the timing of water application, (2) increasing the efficiency of water use by the plants, and (3) increasing the drought tolerance of the plants. The first method is practiced most because it depends on engineering and minimally on the crop. Transporting water with low evaporative loss, preventing runoff, storing water in catchments, delivering water only to the root zone, and timing irrigation to the needs of the plant have been successful in improving productivity per unit of water delivered to the farm. There are estimates that just by improving irrigation timing, the amount of applied water can be decreased by half in some crops while maintaining high levels of production (e.g., Bordovsky et al., 1974). The second and third methods depend on understanding the biology of the crop and whether it can be manipulated to achieve the same productivity with less water. The state of knowledge in this area is the focus of this review.
11. WATER USE EFFICIENCY Water use efficiency (WUE) usually is defined as the total dry matter produced by plants per unit of water used, WUE
D W ’
= -
where D is the mass of dry matter produced (usually aboveground) and W is the mass of water used (usually including direct evaporation from the soil). For a field experiment, D and W would be expressed on the basis of land area. For a single plant, D and W would be measured in the same plant and expressed on the basis of the whole plant. Sometimes, the D is the economically valuable part of the crop (for example the grain, tuber, or fruit) and WUE refers to yield. One may also consider the water use efficiency of a single leaf, and so on. The higher the production per unit of water use, the higher the efficiency. There is extensive evidence that WUE varies among species in the same environment and among climates for the same crop (Briggs and Shantz, 1914; de
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Wit, 1958; Hanks in Taylor er al., 1983; Tanner and Sinclair in Taylor et al., 1983). Taking advantage of the species and climate effects can help manage limited water supplies in agriculture. For example, alfalfa (Medicago sativa L.) has a lower water use efficiency than maize (Zea mays L.) when grown in nearby sites in the same year (Hanks in Taylor et al., 1983; Miller, 1938). Thus, simply by changing crops, water consumption can be reduced with little sacrifice in dry matter production. Relocating production to a new climate with lower evapotranspiration is another possible approach. For economic reasons, however, these options are not often employed and probably will not be until the cost of water rises to a level that forces change. What then are the prospects for improving water use efficiency within a species, or protecting against yield loss in a particular climate when irrigation is not possible?
A. PHYSIOLOGY OF WATER USEEFFICIENCY The dry mass of plants consists mostly of the C and 0 atoms fixed photosynthetically from CO,. These elements are much heavier than the H atoms that also are prevalent in the dry mass and that originate in the water photolyzed during the photosynthetic process. As a consequence, D of Eq. ( I ) represents mostly the net C0,-fixing activity of the plant. Before fixation, the CO, must diffuse into the leaf and dissolve in the wet surface of the cells where it becomes available to be fixed. The wet surfaces are exposed to the atmosphere inside the leaf and transpiration is inevitable. As a result, the photosynthetic cells dehydrate to varying degrees. Water absorbed from the soil replenishes the water lost by the cells, but the water potential of the cells must be low enough to maintain absorption. The stomata and waxy cuticle of the epidermis control the transpiration rate and thus the amount of water needing to be absorbed, and lower water potentials cause stomata1 closure. This regulation of transpiration and absorption affects the balance between net CO, gain and water loss and in turn the WUE. Depending on the leaf anatomy and physiology, the dry matter produced per unit of water used can vary widely. In addition to these factors, water use also is affected by physical factors. CO, enters by diffusing down a concentration gradient to the leaf interior, and the water vapor in the intercellular spaces inside the leaf likewise diffuses in the opposite direction. The lower the external humidity, the faster transpiration will be when all other factors are constant. Leaf temperature plays an important role by affecting the vapor pressure of water in the leaf. The higher the leaf temperature, the higher the vapor pressure and the more rapid the transpiration. Water use will differ among sites and seasons for these reasons and the water use efficiency in Eq. ( I ) thus reflects a complex of plant and environmental factors. Briggs and Shantz (1914) conducted an extensive survey of the water use
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efficiency of crops, and they expressed it as the water requirement, that is, the amount of water transpired to produce a unit Df aboveground dry matter, which is the reciprocal of the water use efficiency. They grew the plants in large containers of soil and made measurements for the entire growing season. This had the advantage that a large number of crops could be compared in a uniform climate during a single season. In their experiments, the transpiration ratio of maize, sorghum (Sorghum bicolor L.), and millet (Panicurn rniliaceum L. and Sefaria italica (L.) Pal. = Chaetochloa) was less than for the other crops and, although Briggs and his co-workers could not have known at the time, the three crops are C , species possessing a special anatomy and biochemistry that allows CO, to be concentrated around the site of fixation. This resulted in more photosynthesis per unit of water transpired and accounted for the lower water requirement. After the experiments of Briggs and his co-workers, various investigators measured water use efficiency under field conditions where all the adaptations of the crop could express themselves (de Wit, 1958; Hanks in Taylor et a l . , 1983). Typically, the experiments involved season-long exposure to differing amounts of irrigation. Figure 1 shows examples for Logan, Utah, where wheat (Triticum aestivum L.) and maize were grown with varying amounts of irrigation in 1975
Wafer Use Efficiency
EVAPOTRANSPIRATION (xi O6 Kg H2O.ha -’)
Figure 1 Production of aboveground shoot dry matter at various levels of water use in several crops near Logan, Utah. The years in which the crops were grown are given in the symbol key. Water use was controlled by irrigation that held conditions essentially constant for the growing season. Water use is the combined evaporation from the soil and transpiration from the plants. A positive evapotranspiration intercept indicates the amount of water obtained from soil stores. The slope of the linear relation is the water use eficiency, which was 4.49 g of dry mass per kilogram of H,O for maize, 2.50 for wheat, 2.36 for alfalfa, and 2.1 I for barley. Maize is a C, plant and the others are C,. Adapted from Hanks (in Taylor et al., 1983).
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and alfalfa and barley were grown in other years (Hanks in Taylor et a / ., 1983). There is a linear relationship between water use and dry matter production. The linearity is mostly caused by the diffusion link between photosynthesis and transpiration because the visible radiation input is almost completely absorbed by crops after the canopy closes and, in a given climate, the input tends to be partitioned in a constant proportion between energy for transpiration and energy for photosynthesis. The slope of the linear relationship is the water use efficiency [Eq. ( I ) ] , and the straight line indicates that the water use efficiency does not vary as the availability of water varies. However, it differs among species, as is apparent in Fig. 1 for the C, maize and C, wheat, alfalfa, and barley. These experiments confirm the differences noted by Briggs and his co-workers and further indicate that water use efficiency does not differ under varying availabilities of soil water in this type of experiment. However, it differs among species, climates, and from year to year (Briggs and Shantz, 1914; Brown and Simmons, 1979; Garrity et al., 1982; Hanks in Taylor et al., 1983; Kawamitsu el al., 1987; Pandey et a / . , 1984a,b; Robichaux and Pearcy, 1984; Tanner and Sinclair in Taylor et al., 1983), and there is a possibility that it will vary with different mineral nutrient availabilities, plant spacing, and other cropping practices. In this respect, it is important to note that while differences between C, and C, species are apparent, similar tests have not been made in species exhibiting Crassulacean acid metabolism. Pineapple (Ananas cornoms (L.) Merr.) has this form of metabolism and it concentrates CO, by temporarily fixing the gas in organic acids at night and releasing it the next day for photosynthesis. During release, the stomata are closed and water is conserved. This allows CAM plants to achieve even higher water saving than C, plants, and limited estimates of water use efficiency are about 20 g of aboveground dry mass per kilogram of water for pineapple (Joshi et a/., 1965; Neales er a / ., I968), compared to 3 to 5 for C4 plants, and 2 to 3 for C3 plants. An alternate approach to the usual description of water use efficiency is to normalize water use for evaporative demand (de Wit, 1958; Tanner and Sinclair in Taylor el al., 1983) and dry mass for the potential productivity of the crop (Hanks in Taylor et a / . , 1983). Thus, modified expressions of WUE have been used, such as
where the fractional dry mass is DID,,,, and is expressed relative to the maximum dry mass produced with optimum water D,,. The fractional water use WI W,,,,, is likewise expressed relative to the maximum evapotranspiration W,,, that would occur with optimum water. This normalization procedure has the advantage that for a water use of, say, half the maximum evapotranspiration, half
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1
0
1
Sorghum
2
4
6
EVAPOTRANSPIRATION(x106 Kg H2O.ha -’) Figure 2 Seed yield at various levels of water use by three sorghum genotypes in West Central Nebraska. The experiment was conducted under conditions similar to those of Fig. 1. The water use efficiency for seed yield is the slope of the line and was 1.8 g of dry mass per kilogram of H,O in RS626. 1.9 in NC+55X, and 1.2 in NB505. The water use efficiency for total shoot dry mass was 3.3 in RS626, 3.2 in NC+55X, and 2.0 in NB505. Adapted from Garrity er al. (1982).
the maximum dry mass would be predicted. This can simplify the job of predicting the impact of water shortages but it requires a knowledge of the maximum dry matter yield and evapotranspiration of the crop for the year, which will vary. For practical purposes, the maximum yield and water use usually are not known and normalization may not be done easily, so the absolute expression in Eq. (1) is preferable. Moreover, farm income for an irrigated crop is generally based on the absolute dry mass or economic yield rather than normalized yields, and expense is based on the absolute amounts of water used. There needs to be a high absolute production of dry mass to justify pumping large amounts of water and there should be a high production of marketable yield. For example, Fig. 2 shows that the water use efficiency differed for production of grain dry mass in sorghum genotypes RS626 and NB505 (Garrity el al., 1982). Normalizing according to Eq. (2) would not distinguish which genotype gives the highest grain production, but Eq. (1) would detect the difference.
B. IMPORTANCE OF HARVEST INDEX The fraction of the crop that is economically valuable, termed the harvest index, is part of the total dry mass and thus part of WUE. There has been a
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general increase in yields of modern crops with little change in the total aboveground biomass, and according to Gifford (1986) the increase is attributable to an increase in the harvest index. The increase has come without much change in the amount of water used, and the result has been a natural improvement in WUE for yield (Richards et al., 1993). There is a maximum to which the harvest index can be increased and the maximum probably is being approached in many modern crops. Therefore, the maintenance of the harvest index is of critical importance when water is in short supply. In Fig. 2, the harvest index was nearly constant among the various treatments. Instead, the differences in WUE were attributable mostly to differences in the total aboveground dry matter (Garrity et a l . , 1982). The experiments involved season-long steady exposure to limited water, and the crops acclimated by growing smaller shoots, with flowering and grain fill adjusting in proportion. In this steady environment, the acclimation allowed the harvest index to be maintained. Under many field conditions, however, plants encounter variable water deficits that do not allow the acclimation possible in long-term experiments, and the harvest index can decrease. This effect can be extreme, and water deficits can give a harvest index as low as zero (see Boyle et a l . , 1991b, for an example). Therefore, considerable opportunity exists for maintaining the harvest index in the face of variable environments, which we will explore more fully later.
C. MEASURING WATERUSEEFFICIENCY The most accurate means of measuring water use efficiency is to monitor the evapotranspiration and harvest the crop for biomass measurements at the end of the season. The WUE can be determined for the total biomass or any part of the biomass. However, these are labor intensive and costly measurements. Less expensive methods have been sought, and one has been to measure directly the COz and H,O exchange of individual leaves (Bierhuizen and Slatyer, 1965; Brown and Simmons, 1979; Robichaux and Pearcy, 1984). Because the CO, molecule contributes most of the dry mass, the gas exchange efficiency can be defined as the ratio of the mass of CO, gained to the mass of H,O lost. Martin and Thorstenson (1988) compared the gas exchange efficiency with the actual water use efficiency for the whole growing season in tomato (Lycopersicon esculentum Mill. ), its wild relative Lycopersicon pentzeflii (Cor.) D’Arcy, and hybrids between them. The relationship was poor because of additional factors affecting dry mass accumulation but not gas exchange. For example, the mass of the plant is determined not only by photosynthesis but also by respiratory losses at night and partitioning to nonphotosynthetic organs such as roots. It is altered by temperature and the molecular composition of the dry mass. Gas exchange for short times during the day does not detect these additional factors. Therefore,
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while the gas exchange efficiency gives valuable insight into the physiologic and metabolic controls that might operate during photosynthesis and transpiration, the method is being used less frequently than in the past. Another method is based on the relative abundance of natural isotopes in plant tissue. Although most of the CO, in the atmosphere is T O , , a small amount is T O 2 . Because the I2CO2 is lighter, it diffuses more rapidly than 13C02. Also, ribulose- 1,5-bisphosphate carboxylase fixes the ' T O 2more rapidly than I3CO2. Consequently, the cells accumulate relatively more 12C than 13C, and the unused C diffuses out according to the extent of stomata1 opening. This outward diffusion is correlated with transpiration. Because the inward diffusion and use of I2CO2correlates with photosynthesis and dry mass but the outward diffusion of 13C0, correlates with transpiration, the relative uptake of I2C and I3C correlates with the water use efficiency. Generally, higher water use efficiency correlates with lower tissue 12C relative to 13C for wheat, peanut (Arachis hypogaea L.), barley (Hordeum vulgare L.), and other crops as shown in Fig. 3. Therefore, the measurements detect differences in WUE among individuals within a species and they only require the ratio of the isotopes in tissue samples to be compared to a standard (Bowman et al., 1989; Brugnoli et al. 1988; Condon et al., 1987, 1990; Farquhar and Richards, 1984; Hubick and Farquhar, 1989; Hubick et al., 1986). The ratio technique makes it possible to survey a large number of plants at moderate cost. Differences integrate the conditions over which the plants were
Discrimination (Woo)
Figure 3 Water use efficiency and carbon isotope discrimination compared in various genotypes of (A) wheat, (B) barley, (C) peanut, and (D) wheatgrass. Adapted from Farquhar and Richards (1984), Hubick ef al. (1988). Johnson et at. (1990). Hubick and Farquhar (19891, and 'hrner (1993).
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grown. Analyzing the entire shoot indicates the water use efficiency for the time required to grow the shoot whereas analyzing only leaf starch indicates the water use efficiency during the time necessary to accumulate the starch. One may integrate over long or short times which avoids one of the problems of the gas exchange technique. Martin and Thorstenson (1988) used this technique to show that differences in water use efficiency were present between the domestic tomato species and L . pennellii and their hybrids. Differences in water use efficiency were detectable in isotope ratio data between the parents and the hybrids particularly when water was optimally available. The domesticated parent had the lowest efficiency, the wild parent the highest efficiency, and the hybrids showed intermediate behavior. Because the species could be crossed, it was possible to correlate the differences in water use efficiency with restriction fragment maps of the tomato DNA (Martin et af.,1989). Three loci were found to be predictors of the variation in water use efficiency in field grown tomato. This landmark effort indicates that water use efficiencies are determined by relatively few genetic loci and implies not only that agriculturally relevant differences exist but that they might be genetically manipulated in a simple fashion. The success of the method suggests that differences in water use efficiency exist in individual species and might be usefully incorporated into breeding programs, although this is still in its infancy (Bowman et al., 1989; Brugnoli et al., 1988; Condon et a l . , 1987, 1990; Hubick and Farquhar, 1989; Hubick e t a l . , 1986). Genetic variation clearly exists but in crop canopies the variation becomes less clear. Substantial water limitation usually gives a negative relationship between discrimination and WUE as shown in Fig. 3, but under relatively favorable conditions, the relationship tends to become less negative or even positive. In the field, this can obscure relationships developed from pot experiments. For example, Condon and Richards (1993) showed that two wheat genotypes differing by 40-50% in leaf diffusive conductance only differed by 15% in canopy transpiration efficiency. In the field, this difference was not reflected in improved WUE for the crop because soil evaporation differed in opposition, canceling the 15% effect. There was a slower development of the canopy in one genotype than in the other, and this was responsible for the canceling effect (Condon and Richards, 1993). Therefore, differences in WUE may need to be combined with other crop traits to be realized as water savings.
D. MOLECULAR BIOLOGYAND WATERUSEEFFICIENCY Opportunities to improve WUE generally involve many genes and many interactions. Perhaps this is not surprising in view of the massive changes in plant form and anatomy that have occurred as plants colonized the land. The develop-
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ment of roots, cuticle, stomata, vascular systems, and seeds all occurred after plants left their original water environment and, in evolutionary time, these developments are relatively recent. As a consequence, single genes that substantially change the WUE of plants are difficult to find. With the increased ability to transform plants genetically in recent years, it would be desirable to apply the tools of molecular biology to the improvement of WUE. However, the tools work best when single genes are manipulated. The problem can be seen in C, species where many genes code for enzymes and leaf anatomy that differ from those in C, species. While C, photosynthesis confers a clear improvement in WUE, the genetic complexity of the trait makes it difficult to incorporate in C, plants. However, there are traits of more limited scope that probably have less complex genetics and might benefit from the application of molecular biology. Some of these are considered below as ways to improve the drought tolerance of the plant.
111. DROUGHT TOLERANCE Plants showing improved growth with limited water are considered to tolerate drought regardless of how the improvement occurs or whether the water use efficiency is affected. Some species can avoid drought by maturing rapidly before the onset of dry conditions or by reproducing only after rain. Examples of these drought avoiders are ephemerals such as California poppy (Eschscholtzia californica (Cham.)) that can complete their life cycle in a few weeks, or tree crops such as coffee (Cofea arabica (L.)) and cacao (Theobroma cacao (L.)) that flower and fruit after drought followed by rain (Alvim, 1960, 1985). Others can postpone dehydration by growing deep roots or sealing themselves tightly against transpiration or accumulating large stores of water in fleshy tissues. Examples of dehydration postponers are upland rice (Oryza sariva L.) with deep roots compared to paddy rice (Chang et al., 1974), or agave (Agave deserti (Engelm.)) or saguaro cactus (Carnegiea gigantea (Engelm) Britt. and Rose) with thick cuticle or fleshy tissue. Still other species allow dehydration of the tissues and simply tolerate it by continuing to grow when dehydrated or by surviving severe desiccation. Certain intertidal algae such as Fucus vesiculosus (L.) or lower vascular plants such as Selaginella lepidophylla (Hook. & Grev.) can carry out photosynthesis at very low water contents and tolerate desiccation to the air dry state without losing viability. The seeds of most angiosperms also can tolerate severe desiccation. These effects are generally distinct from the factors controlling water use efficiency. Drought avoiders depend on the timing of development which is under internal control. They tend to reproduce themselves after a minimal accumulation of dry matter and their success ensures that they are represented in the next
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generation. Dehydration postponers having deep roots may have a water use efficiency identical to that of other species but will accumulate more dry weight because of their ability to gain access to a larger amount of water than shallow rooted types. In effect the slope of the water use efficiency relation in Fig. 4 may be the same but the deep rooted species work farther out on the curve. Their adaptations are mostly structural and take time to build, requiring the expenditure of photosynthetic products. Finally, dehydration tolerators may have the same water use efficiency as dehydration-sensitive species when water is available but the tolerators can grow at tissue hydration levels that the other species cannot. Of the three forms of drought tolerance, dehydration tolerance is most intriguing because it often requires only slight repartitioning of dry mass. An example is osmotic adjustment (Morgan, 1984; Munns, 1988) which occurs because dry mass normally used to synthesize new cells instead accumulates in the cells as solute (Meyer and Boyer, 1972; 1981) or is deposited in fewer or smaller cells (Fraser et al., 1990; Sharp et a l . , 1990). Only a brief decrease in biosynthesis of tissue is necessary to accomplish this (Meyer and Boyer, 1981), but the increased concentration of solutes can markedly increase the ability of the cells to extract water from the soil. The increased solute is present only under dry conditions. In other words, there is little cost to the plant when water is scarce and no cost when water is plentiful.
A. APPROACHESTO IMPROVING DROUGHT TOLERANCE From these examples it can be seen that water use efficiency is important, but crop improvement under conditions of limited water involves more than water
-
EVAPOTRANSPIRATION (x106 Kg HpO.ha -')
Figure 4 Effect of increasing the amount of water available to a crop without changing the water use efticiency. Production moves from A to B. An example might be increasing rooting depth.
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use efficiency. Basing improvement solely on water use efficiency is tempting because breeding programs could select solely for high productivity when water is plentiful. The idea is that, for a given climate, water use efficiency will be highest when dry matter production is highest and the linear relationship of dry matter to water use (Figs. 1 and 2) would allow the high productivity to carry over to drought conditions. However, it is clear that many opportunities will be missed if superior selections are based only on this concept of water use efficiency. Characters such as osmotic adjustment are called into play only during a water deficit. Roots may penetrate deeper soil layers or leaves may persist better during a water deficit in some genotypes than in others, and so on. Without plant selection under water deficient conditions, these beneficial traits will be missed.
800C Morrow Plots
Urbana, Illinois Continuous Maize
0
600C
m
A
5n
5
-
-0 Q
First
E al
n 4000
E
e U *9 0
.i?
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2000
0
1
I
I
I
1890
1900
1910
1920
I
1930 Year
I
I
I
1940
1950
1960
l! '0
Figure 5 Yields of continuous maize crops at Morrow Plots at University of Illinois. The first crop was planted in native prairie. Soil nutrients and organic matter were added to some of the plots beginning around 1905. Hybrids were introduced to all plots in 1937. Yields are 5-year averages. Genotypes are those popular at the time.
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An example of the failure of the high dry-matter approach is indicated by the experience with hybrid maize. Its introduction in the mid 1930s in the United States resulted in immediate increases in yields partly because of heterosis which gave increased dry mass accumulation and because of improved harvest index which gave more grain. However, in soils of low nutrient and water holding capacity, these benefits were not realized. Figure 5 shows that maize yielded moderately well in the Morrow Plots at the University of Illinois in the years 1885-1900 when the prairie soil was first cultivated. The yield became low by 1900 and remained low for many years if the crops received no added nutrients because the soils became depleted in nutrients and organic matter and became drought prone. With the introduction of hybrid maize in 1937, heterosis caused a dramatic increase in the dry matter production if nutrients were available but there was no yield increase in the impoverished plants (Fig. 5). The capability for high grain yield was expressed only later in the 1950s when hybrids were introduced that could be planted at higher densities, which increased nutrient and water acquisition because of the greater number of plants. Thus, selections for high productivity under favorable conditions did not improve productivity under unfavorable conditions. Other features of the crop had to change to allow the crop to cope with environmental limitations before increased productivity could be realized.
B. EXAMPLES OF IMPROVED DROUGHT TOLERANCE 1. Relation to Yield under Favorable Conditions The design of a crop improvement program for drought tolerance seems difficult at first because water is so ubiquitously involved in growth and metabolism that identifying targets seems impossible. Moreover, the multitude of possible targets implies that effects might involve enormous numbers of genes, and improvements might be only incremental or, worse still, might cause problems at other genetic loci. However, there are examples of successful approaches that have resulted in significant improvements in the drought tolerance of plants. Jensen and Cavalieri (1983) developed improved drought tolerance in grain production of maize by making promising crosses and testing at varying water availability in the field at a large number of locations. Genotypes were identified that had all the combinations of yield performance under optimum and waterdeficient conditions: superior yield under both conditions, high yield under optimum conditions but low yield under water deficit, and low yield under optimum conditions but high yield under deficit. The first kind of response is the preferred one but the last response seems worthy of some consideration. The study by Jensen and Cavalieri (1983) is particularly important because it
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tested whether improved yield under water-limited conditions sacrificed yield under optimum conditions. Grain yield was determined from about 500 field replications, and whole season yield performance was determined to give a wide test of plant performance and to allow characters associated with particular environments or parts of the life cycle to be expressed. Because of the replications and experimental design, their experiments give the strongest possible evidence that improvement under water-limited conditions need not sacrifice yield under favorable conditions. Sojka et al. (1981) found a similar result in wheat. The principle was confirmed by Morgan (1983) in a completely different experiment. He selected wheat for superior osmotic adjustment under dehydrating conditions and observed improved yields that were at no cost to yield under optimum conditions. An important feature of this study was that the test genotypes had essentially the same genetic background. Therefore, osmotic adjustment was the main difference between the standard commercial genotype and the genotype with superior drought performance. Quisenbeny et al. (1984) also tested the effectiveness of osmotic adjustment in cotton (Gossypium hirsutum L.) but concluded that it had little benefit. However, they failed to compare plants of similar genetic backgrounds and the results could have been caused by features other than osmotic adjustment. Therefore, the work of Morgan (1983) provides the better test and indicates that there can be a benefit of osmotic adjustment without sacrificing yield under optimum conditions. This probably is explained by the low metabolic cost of osmotic adjustment together with the lack of osmotic adjustment under optimum conditions.
2. Physiological Correlations The success of Jensen and Cavalieri (1983), Sojka el al. (19811, and Morgan (1983) illustrates that it is useful to know the physiological attributes that give superior yield under water limited conditions. Blum et al. (1989) explored these attributes in sorghum where various genotypes were grown with little or no rain during the growing season so they were forced to use stored soil water (Fig. 6). As Jensen and Cavalieri (1983) and Sojka ef al. (1981) observed in maize and wheat, the range of genotypes displayed several combinations of performance in the two environments. For example, genotypes ICSV-193, ICSV-I 12, RS-610, ICSV-186, and ICSV-I had high yield when water was made available and yielded well in the water-limited environment. Genotype ICSV- 1 had especially high productivity when water was available. ICSV-110 did poorly in both environments. The physiological attribute most closely associated with success was early heading. Water was being steadily depleted from the soil profile, and early heading favored grain growth while water was still available. The successful genotypes also displayed higher harvest index, higher leaf water potentials, and
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-
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Genotype Figure 6 Grain yield of various sorghum genotypes when grown without rainfall under irrigated and nonirrigated conditions. Yields without irrigation are shown as stress yields and as a percentage of the yields with irrigation. Adapted from Blum ~t ol. (1989).
cooler leaves, presumably reflecting superior water acquisition compared to the poorer performers. This work is particularly valuable because it identifies traits that might be used to accelerate selections for improved drought performance. Blum and his co-workers (1991, 1992) also compared the performance of sorghum hybrids, open pollinated varieties, and landraces from local farmers. During water deficits, better performance was observed in the hybrids having the higher harvest index and in the superior varieties having larger aboveground dry mass. The landraces had generally low performance. This suggests that further improvement might be possible if higher harvest index, larger dry mass, and early flowering could be combined in sorghum growing mostly on stored soil water. The control of the life cycle for earliness also was valuable in cowpeas (Vigna unguicitlutu (L.) Walp.). Hall and Grantz ( I98 I ) selected early flowering cowpeas that escaped late season drought. Because the reproductive tissues were the valuable structures, genetic selection for earliness restricted growth to the part of the season when water was available. Similarly, Passioura (1972) demonstrated that wheat produced grain if the roots were pruned to reduce early season water use but not if the roots were unpruned and the plants consumed most of the soil water before grain fill. Richards and Passioura (1981a.b) used genetic means to incorporate this conservative water use into commercial wheat genotypes. It was
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possible to cross commercial genotypes with landraces having a smaller vessel diameter in the root xylem. The decreased vessel diameter was heritable and field trials indicated that the modified genotypes had slightly improved grain yield in fields with late-season drought in which the crops grew mostly on stored soil water (Richards and Passioura, 1989). However, there was a yield penalty in the modified genotypes when water supplies were adequate because the landrace bearing the diameter trait had lower yield characteristics than the commercial lines.
3. Importance of Deep Rooting Deep rooting probably accounts for a major share of the differences in drought tolerance between species. Trees can remain vigorous during droughts that completely inhibit the growth of grasses and other smaller plants mostly because tree roots can extend to deeper water-containing soil layers than the roots of smaller plants. Maize and sorghum roots grow to depths of 2 to 3 m and the plants grow and remain green when surrounding short grasses with shallow roots become brown due to soil dehydration. Taylor and Terrell(l982) give a detailed list of the depth and extent of rooting of various species. Burton et al. (1954, 1957) showed that deep-rooted Bermuda grass (Cynodon dactylon (L.) Pers.) exhibited increased pasture productivity compared to more shallow rooted types in humid regions subjected to sporadic drought. Upland rice yielded better than paddy rice in upland environments (Chang et al., 1974) because the upland rice had deeper roots. Deep roots show the greatest benefit in deep soils that allow deep rooting to be expressed, although there also can be an advantage to extensive rooting in shallow soils. There are genetic differences in the ability of roots to penetrate deep soil layers (Boyer et al., 1980; Hurd, 1974; Jordan et al., 1979; O’Toole and Bland, 1987; Taylor et al., 1978). In some instances, the differences are inherited simply. Ekanayake et al. (1985) and Armenta-Soto et al. (1983) showed that in rice the difference in depth of rooting was controlled by only a few genes. Rice extends from deepwater genotypes that are semiaquatic to paddy rices with shallow roots and finally to upland rices that require well-drained soils and are deeply rooted (Chang et al., 1974). Thus, wide genetic variation for rooting exists in this crop. In wheat, Hurd (1974) found large genetic differences in deep rooting that led to improved cultivars. Boyer et a / . (1980) found evidence that the high yields of modern soybean cultivars were traceable in part to less midday dehydration of the leaves resulting from deeper rooting than in older cultivars. Frederick et al. (1990) did not find differences in midday dehydration of some of the cultivars, but their measurements of water status involved excised tissue stored for considerable time, which may have obscured the differences. O’Toole and Bland (1987) provide a thorough and useful review of root differ-
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ences and genetics in crops and point out that the benefits of deep rooting could be exploited much more than in the past. 4. Importance of Cuticle
The cuticle is a waxy layer covering the outer surfaces of the entire shoot. Waxes on leaves cover the epidermis and guard cells of the stomata and extend into the interior leaf surfaces where they become thin (Boyer, 1985; Norris and Bukovac, 1968). When stomata close, water loss is controlled by the cuticle and any remaining vapor diffusion through the stomata. Fruits and stems often do not contain stomata, and water loss is controlled entirely by the cuticle. There are genetic differences in leaf water loss when stomata are closed. Among the species soybean, cowpea, pigeonpea (Cajanus cajan (L.) Millsp.), and black gram (Vigna murzgo (L.) Hepper), Sinclair and Ludlow (1986) attributed differences in survival during severe dehydration to differences in epidermal conductance when stomata were closed. Within species, differences have similarly been found in the dark in wheat (Rawson and Clarke, 1988) or during water deficits in soybean (Glycine mux (L.) Merr.), wheat, oat (Avena sarivu (L.)), and maize (Bengston et ul., 1978; Dub6 et al., 1975; Paje et al., 1988; Rawson and Clarke, 1988). It has been proposed that this variation could be used to select for genotypes having decreased conductance when stomata are closed and that the water saving could be substantial (Jordan et a l . , 1984; Rawson and Clarke, 1988; Sinclair and Ludlow, 1986). The variation has been studied mostly in crop plants having stomata on both leaf surfaces and some of the variation could thus be caused either by incomplete stomata1 closure or by differences in cuticular wax development. The exact contribution of each is difficult to evaluate; however, Jordan et al. (1983, 1984) showed that there is genetic variation for epicuticular waxes in sorghum and Jenks cr a / . (1994) described sorghum mutants differing in amounts of cuticular wax. In barley, there are mutants that differ in amounts and composition of cuticular waxes (von Wettstein-Knowles, 1989). Similar variation was also observed in cuticles of oat (Bengston el al., 1978) and rice (O’Toole et a l . , 1979). The amount of wax in rice was markedly less than that in barley, wheat, and sorghum, and was less in paddy rice than in upland rice (O’Toole, 1982; O’Toole et ul., 1979). Dry atmospheric conditions can desiccate reproductive structures in rice probably because of the thin cuticle, and abortion results (O’Toole et ul., 1984). This suggests that significant improvement in dryland performance might be possible in rice if the cuticle contained additional wax. An example of genetic improvement of drought tolerance based on epicuticular wax deposition is the selection for improved seedling establishment in native range grasses in the western United States. Wright and Jordan (1970) showed rapid improvement in the establishment of boer lovegrass (Eragrostis curvulu
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Nees) selected for seedling growth in dehydrated soil. In this situation, stomata1 conductance was relatively unimportant and the character that appeared most improved was the thickness of the cuticle covering the shoot tissues of the young seedlings (Hull et al., 1978). These selections allowed the establishment of grasses to become more reliable when rooting was shallow, rainfall was sporadic, and germination had to occur with limited water.
IV. WATER DEFICITS AND REPRODUCTION Reproductive development holds particular interest for improving drought tolerance in crops because a large part of agricultural production is devoted to the reproductive parts of plants. in the United States, reproductive crops (grain, fruit, nut, vegetable) account for about 78% of the harvested area of land. Moreover, early stages of reproduction are more susceptible to losses from limited water than any other stage of development in reproductive crops (Claassen and Shaw, 1970; Salter and Goode, 1967). A good example is maize, in which part of the problem is caused by a high susceptibility of floral parts to inhibition of cell enlargement (Herrero and Johnson, 1981; Westgate and Boyer, 1985b). This susceptibility exists in part because the cells enlarge dramatically in the floral tissues during normal development, and water deficit can prevent the enlargement. However, more than cell enlargement is involved because Damptey et al. (1978) observed losses in inflorescence development in maize treated with abscisic acid during floral initiation before most enlargement of reproductive structures had begun. Losses in reproductive activity also were reported because of megagametophyte sterility (Moss and Downey, 197l), asynchronous floral development (Herrero and Johnson, 1981), and nonreceptive silks (Bassetti and Westgate, 1993; Lonnquist and Jugenheimer, 1943), depending on when dehydration occurred. When gamete and floral development are normal and plants are hand-pollinated, reproductive failure still occurs and can be induced by only a few days of dehydration (Westgate and Boyer, 1985b, 1986b). The loss is caused by irreversibly arrested embryo development (Westgate and Boyer, 1986b). This indicates that, provided cell enlargement allows for good floral development, there can be good pollination and fertilization even when water deficits exist. Nevertheless, a complete block in embryo growth may remain when everything has been normal up to the time of cell division in the newly formed zygotes. In other crops such as wheat and barley, drought during microsporogenesis caused pollen sterility (Morgan, 1980; Saini and Aspinall, 1981; 1982; Saini et a / . , 1984). Well-watered plants whose stems were fed abscisic acid (ABA; Saini et al., 1984) or whose shoots were sprayed with ABA (Morgan, 1980) showed a
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similar pollen abortion, thus implicating high ABA levels during dehydrating conditions. However, the high ABA may have acted by closing stomata and inhibiting photosynthesis. Increasing CO, pressures around wheat plants overcame some of the reproductive losses (Gifford, 1979), which supports an involvement of photosynthesis. In rice, dehydration of the soil caused especially severe dehydration of reproductive tissues, and death and bleaching of florets followed probably because of inadequate cuticular wax (O’Toole et al., 1984). Therefore, in various crops, there is increasing evidence for metabolic .and growth regulator effects and some direct dehydration effects that might account for the susceptibility of early reproduction to water limitation. CO, and ABA seem to be involved, and photosynthesis also may play a role but each could act in concert or separately, depending on the crop. Additional insight may be possible from studies of embryo development in maize. Westgate and Boyer (1985a) found that the block in embryo development was correlated with low photosynthetic reserves in the maternal plant. Because photosynthesis was inhibited during the treatment, the lack of reserves could have caused embryo starvation. Westgate and Thomson Grant ( 1989) observed that the sugar content of maize embryos was not significantly different in hydrated and dehydrated plants but concluded that the flux of sugar might differ. Schussler and Westgate (1991a,b) found that the uptake of sugars was less in maize ovules isolated from dehydrated plants even though the sugar content was high, which further confirms that the flux of sugars was more important than the sugar content of the developing grain. Myers et ul. (1990) showed an inhibition of endosperm cell division in maize when high ABA levels were present 5 to 10 days after fertilization.
A. MECHANISMS OF LOSSES Boyle et a / . (1991a,b) took advantage of the finding of Westgate and Boyer (1985a; 1986a) that a few days of low water potentials can prevent embryo growth and developed a system to feed stems a complete medium for embryo growth during this time. This allowed photosynthetic products and other salts and metabolites to be supplied to the plants at normal levels without rehydrating the plants. The controls yielded well, but withholding water for a few days virtually eliminated grain production because of embryo abortion. Production was almost fully restored when the plants were infused with the complete medium as low water potentials developed. Infusing the same amount of water alone showed no restorative activity. Therefore, it was possible to maintain reproduction by feeding substances normally supplied by the parent plant during embryo development, which indicates that sufficient water was available to the embryos
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so that water itself was not the limiting factor, and embryo growth had been blocked by some other substance(s) that the parent plant failed to supply. Thus, reproductive loss appears to be a biochemical problem. Zinselmeier et al. ( I 995a) found that the active ingredient was sucrose, and ovary sucrose had to be elevated above controls in order for embryo abortion to be prevented. Because sucrose is the main translocation form of photosynthate in maize, this finding showed not only that the lack of photosynthate was the basis of abortion but also that there was a block in the utilization of sucrose by the ovaries (Zinselmeier et al., 1995b). Abortion was accompanied by a loss of starch, which is a product of sucrose metabolism in the ovaries, and starch returned partially toward control levels when sucrose was fed. This suggests that the block could have been located between the sucrose supply to the ovaries and starch synthesis in the ovaries, although there is a possibility that starch degradation also was involved. It is worth noting that little is known about the role of ovary starch. In contrast to endosperm starch, which is a terminal pool and has received a great deal of attention, ovary starch is not a terminal pool. It forms before fertilization and, because it decreases during dehydration and recovers when sucrose is fed to the stems of dehydrated plants (Zinselmeier et al., 1995a), it appears to be mobilizable. Apparently, under unfavorable conditions for the parent plant, the products of breakdown are used to support the growth of the ovary tissues. Fader and Koller (1985) suggested that ovary starch could be important for developing soybean pods. Important insight may be gained from a fuller understanding of ovary starch in early reproduction. Edmeades and his co-workers found that the time between pollen shed and silking can be changed by genetic means in maize, and they used early silking to indicate vigorous development of the ear (BolaAos and Edmeades, 1993a,b; Bolafios ef al., 1993; Edmeades et al., 1992, 1993). Early ear development may indicate that the plant supplies more of the biochemical requirements for ear growth and may be a genetic means of accomplishing the same result as feeding sucrose to the stem (Zinselmeier et al., 1995a). In effect, early ear development may be a visual signal for enhanced sucrose availability or utilization by the ear. Fischer et al. (1989) selected a population of tropical maize for several physiological attributes likely to improve drought tolerance (low canopy temperature, low leaf death, early silking relative to pollen shed, days to anthesis) and found that most of these traits conferred a yield gain, but early silking relative to pollen shed generally accounted for more of the yield gain than the other traits. After three cycles of selection, grain yield increased by 320, 420, and 4 10 kg-ha- in the mild, medium, and severe dehydration treatments but not at the expense of yield in hydrated conditions. The authors point out that progress was accelerated by using physiological characters in addition to grain yield in the selection program.
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Tollenaar and Mihajlovic (1991) report that the genetic improvement of maize yields was associated with improved resistance to the herbicide bromoxynil (4hydroxy-3,5-dibromo-benzonitrile), which inhibits photosynthetic electron transport at photosystem I1 and probably other aspects of energy metabolism. Tollenaar et d.( I 994) suggest that the mechanism may be related to the production of oxygen-containing free radicals that may be more rapid during drought or other unfavorable environments. The production of these agents is destructive to membrane components, particularly photosynthetic membranes; further, chloroplasts contain high levels of antioxidants (glutathione, hydroquinones, ascorbate, tocopherols, carotenoids) and the cells possess enzymes (peroxidases, catalases, superoxide dismutases) that probably are protective. Tollenaar et al. ( 1994) express the view that modern maize hybrids may have been selected for improved levels of the protective components that would be expressed in protection against bromoxynil. However, it also seems possible that the selection may have been toward less penetration of the leaf or greater metabolic degradation of bromoxynil. These experiments offer the promise of identifying components that may protect against losses in early phases of reproductive development when plants are subjected to moderate dehydration. Selection for genotypes that store significant amounts of mobilizable photosynthate during early reproduction is one approach. Avoidance of early leaf senescence, which decreases photosynthetic capacity, might be another. Regardless of the approaches taken, it is clear that the reproductive fraction of the plant can vary from zero to nearly normal during a drought, which implies that successful protection of reproductive development may be possible by genetic and cultural means under otherwise inhibiting drought conditions.
V. DESICCATION TOLERANCE When seeds mature, it is common for them to dehydrate as part of the maturation process. Barlow et al. (1980) found water potentials as low as -5MPa in maturing wheat grain. Westgate and Boyer (1986~observed ) water potentials of -7 to -8 MPa or lower in maize grain late in the growing season. These seeds are exposed somewhat to the atmosphere and are known to desiccate to a large extent by evaporation to the air. Seeds surrounded by a fleshy fruit show a similar but less severe desiccation. Welbaum and Bradford (1988) observed that water potentials of melon seeds (Cucurnis melo L.) decreased to about -2 MPa during maturation, and the surrounding fleshy fruit decreased similarly in water potential. Bradford (1994) considers high solute concentrations to be present in the apoplast surrounding embryos and proposes that structures may exist to keep the
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solutes localized there, The low osmotic potential of the apoplast solution may explain how the seeds are dehydrated inside fleshy fruits. Regardless of whether the seeds air-dry or are dehydrated osmotically inside a fruit, it is clear that embryos become exceptionally tolerant of desiccation late in maturation despite their susceptibility to the effects of water limitation when they are young. Plants lower in the evolutionary scale than seed plants sometimes show a similar tolerance to desiccation. Some fungi, algae from the intertidal zone, and a few mosses and lycopods can be desiccated to the air-dry state without losing viability (Bewley, 1979). There also are some specialized seed plants (Craterostigma species, Myrothamnus jiubellifoliu (Welw.), Xerophyta species) that can tolerate desiccation (Gaff, 197I , 1977; Gaff and Churchill, 1976). However, desiccation tolerance is virtually nonexistent in most agricultural species except for the seeds and pollen. It is curious that most seed plants, which are descendants of plants that crossed the intertidal zone, should have lost the ability to tolerate the desiccation that is so prevalent in that zone. In land plants, desiccation tolerance often evolved as part of the seed habit because an aqueous medium generally was absent and the pollen and ultimately the embryo were exposed to drying conditions during dispersal. In agriculture, this property makes it possible to store seeds and allows uniform planting times. However, after germination, the plant generally loses its desiccation tolerance and remains sensitive for the rest of the life cycle until pollen is produced. Pollen can desiccate to a remarkable degree in species such as maize without losing viability (Westgate and Boyer, I986a.b).
A. MOLECULAR FEATURES OF DESICCATION TOLERANCE An important aspect of severe desiccation is that water contents can become so low in the cells that enzyme activities can be directly inhibited by the lack of water, as described by Vertucci and Leopold (1987a,b). Enzymes are affected directly when sufficient water is lost to remove the hydration shells next to the protein. The activities begin to decrease when the monolayer of water next to the peptide surface is all that remains. When the monolayer begins to be lost, activity decreases and disappears when water covers only a few polar sidegroups in the peptide backbone (Rupley et al., 1983). Substrates probably are unable to reach the active site of the enzyme because the aqueous medium is no longer continuous (Skujins and McLaren, 1967). Cells and tissues begin to show these effects when they are desiccated in atmospheric humidities around 60-70% (water activities of 0.6-0.7) and below (Skujins and McLaren, 1967). Thus, seeds desiccated to the air-dry state are likely to be affected by these phenomena. Most can return to activity when they are rehydrated, provided water contents have not
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become so low that the tightly bound water required for viability is lost (Vertucci and Leopold, 1987a,b). On the other hand, leaves generally are susceptible to desiccation damage at much higher water activities. When dehydrated to the air-dry state, leaves of most crop species show a breakdown of compartmentation that releases cell constituents to the apoplast (Leopold et al., 1981), and the plasmalemma and tonoplast show breakage followed by a loss of organelle structure starting at water activities of about 0.98 (Fellows and Boyer, 1978). In leaves of tolerant species, the membranes and organelles remain intact at low activities although they often are distorted (Hallam and Gaff, 1978a,b). Therefore, an important distinction between tolerance and sensitivity to severe desiccation appears to be the maintenance of membrane structure and an ability of enzyme activity to return upon rehydration. It has been proposed from work beginning with desiccation-tolerant animals that a possible mechanism to account for preservation of enzymes and cell structure might be an accumulation of specific sugars such as trehalose or sucrose whose structure resembles water in certain respects (Crowe and Crowe, 1986). Sugars having the appropriate stereostructure might form hydrogen bonds with cell membranes where water would ordinarily bind. Because the sugars would remain as water is removed, the bonding would be stable and membrane structure might be maintained where otherwise it would become disorganized. Evidence that the sugar replacement hypothesis may have merit is the accumulation of sugars such as sucrose and raffinose in developing seeds (Caffrey et al., 1988; Koster and Leopold, 1988). Species such as maize have seeds that can tolerate desiccation to the air-dry state, and their sugar concentration, while not high for the seed as a whole, becomes high in the remaining residual water of the drying seed and could have a stabilizing influence at local sites. As germination proceeds, the stabilizing sugars are metabolized to nonstabilizing ones such as glucose and fructose, and desiccation tolerance is lost (Koster and Leopold, 1988). A related hypothesis is that certain sugars may be converted to the glassy state during dehydration (Williams and Leopold, 1989). The glassy state is common in sugars such as sucrose used to make candy, and evidence for the existence of glassy sugars is accumulating for embryos of dehydrated seeds (Williams and Leopold, 1989). A similar role has been proposed for certain proteins in seeds (Crowe and Crowe, 1986; Dure el a l . , 1989). The developing seeds of a range of crops accumulate hydrophilic proteins in the embryo as normal desiccation begins (Dure et a l . , 1989). The proteins have been variously called dehydrins, embryo maturation (Em) proteins, or late embryogenesis abundant (LEA) proteins (Dure et a l . , 1989). Common to all of them is a high content of hydrophilic amino acids so that the proteins as a whole are highly water soluble. In some of them, an
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alpha-helix is present that could remain structurally stable during desiccation and it has been proposed that this portion of the protein could act like a membranestabilizing sugar (Crowe and Crowe, 1986). The mRNAs for these proteins are not readily detected in leaves or roots of hydrated plants but can be induced by severe desiccation in very young rape (Brassica napus (L.)) (Harada er a l . , 1989) and maize and barley seedlings (Close et al., 1989; Close and Chandler, 1990). There was a marked increase in dehydrin mRNAs when young wheat seedlings were dehydrated soon after germination (Close and Chandler, 1990). The mRNA expression was especially increased in shoots, which are most exposed to dehydration under natural conditions. This cellular response suggests that the dehydrin-Em-LEA proteins play a role in the desiccation tolerance of seedlings. Also, the mRNAs can be induced by treating hydrated seedlings or immature embryos with high abscisic acid concentrations (Galau et al., 1986; Hong et al., 1988; Mundy and Chua, 1988). Abscisic acid levels normally increase in plants subjected to dehydration (e.g., Beardsell and Cohen, 1975) and they become high in maturing dehydrating seeds (Ihle and Dure, 1972). The induction of the mRNAs suggests that there is molecular control that might be manipulated genetically, thus altering the development of desiccation tolerance of young seedlings and embryos.
VI. CONCLUSIONS Land plants appear not to be optimally adapted to water shortages imposed by the environment and indeed we likely would see large improvements in dehydration performance if this chapter could be written after a few hundred million years to give additional time for beneficial adaptations to evolve. Certain metabolic changes have developed during the course of evolution that have improved the ability of plants to withstand limited water supplies, particularly in photosynthesis. The recent evolutionary development of C, photosynthesis and Crassulacean acid metabolism are clear examples, and there is increased water use efficiency in those species possessing these adaptations. Methods of plant breeding and genetic modification can speed the transition to more efficient water use and considerable success has already been achieved. Water acquisition has been improved by deep rooting and strong osmotic adjustment, cuticular characters have been modified to conserve water, and earliness in reproduction has been used to avoid late season droughts. It also appears increasingly possible to improve water use efficiency by genetic means using new techniques for screening for this trait. Water is required for biological activity, and studies show that water use
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efficiency is unchanged by the water supply when the water limitation remains stable for the entire season. This allows acclimation to occur, and the harvest index often remains unchanged as well. However, when the water supply is not stable for the growing season but slowly declines or varies with rainfall, dry matter is partitioned differently to plant parts depending on the time in the life cycle. This is particularly true for reproductive structures and can lead to a variable harvest index, which can be relatively independent of the overall water use efficiency. Attention to the harvest index may provide a means for maintaining the economically valuable parts of a crop even though the total plant dry mass may decline. The demonstration that reproductive losses usually associated with drought may have a biochemical origin raises the possibility that metabolic modifications may be useful for improving the harvest index with limited water, and genetic approaches are being applied to this problem. Also, the molecular mechanisms of desiccation tolerance suggest that changes in expression of specific genes are correlated with decreased lethality of severe desiccation at least during late seed development. From these principles, it is possible to distill certain conclusions that may help in efforts to improve the efficiency of water use and drought tolerance of plants. The approaches at first appear diverse and the complexity makes it tempting to take shortcuts such as selecting seedlings for rapid growth only under favorable conditions or in osmotica, or by using single biochemical tests for performance. In general, the temptation should be avoided because the results have not carried over to field situations. The approaches that have given the most rapid progress in improving drought performance have been: (1) using realistic soil conditions, (2) testing with adequate water and with limited water, (3) understanding the sources of crop failure in the proposed growing area, and (4) targeting a limited number of traits for improvement. In most examples of improvement, there was an intimate knowledge of the soil, climate, and physiology, and biology of the crop. Physiological tests sometimes could be employed to increase the rate of progress, and the problem could be reduced to a few traits to simplify the selection effort. Great progress was made under conditions of realistic water limitation in soils because droughtadaptive factors were called into play and had an opportunity to express themselves. This avoided the problem of selecting only genotypes yielding well in favorable environments that "crashed" in water-limited environments, or using pots that restricted root development and prevented the expression of this important tolerance character. It is now clear that successful improvement of drought performance can come at no sacrifice to performance under favorable conditions but this can be determined only if performance is tested under both favorable and unfavorable conditions.
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THEAFLATOXIN PROBLEM WITH CORNGRAIN Neil W. Widstrom US.Department of Agriculture, Agricultural Research Service, Georgia CoSstal Plain Experiment Station, Tifton, Georgia 3 1793
I. Introduction 11. Background and Identification of Aflatoxins as Contaminants of Corn A. Detection and Quantification of Aflatoxins B. The Scope of Corn Contamination 111. Conditions Impacting AspergilhJuvlls Group Infection and Aflatoxin Accumulation A. The Influence of Factors beyond Grower Control B. Preplanting Considerations n! Managing Conditions during Plant Growth and Ear Development A. Minimizing Stresses on the Corn Plant B. Monitoring Problem Areas V. Handling the Grain Crop at Harvest A. Optimum Timing of Harvest B. Grain Handling and Assessment of Infection and Contamination at Harvest C. Processing the Crop for Market or Home Storage VI. Storage and Utilization of the Final Product A. Monitoring Stored Corn B. Human Consumption and Its Impact on Health C. Contaminated Corn as Animal Feed D. Decontamination Processes E. Use of Contaminated Grain for Ethanol Production F. Contaminated Grain and the Milling Industry VII. Long-Range Solutions A. Breeding Resistant Hybrids B. Interrupting Toxin Production by the Fungus VIII. Conclusions References
2 19 Advances m A p n m y , l'blumr 16 Copyright 0 1996 by Academic Press, Inc. All rights of rrpruductlon in any form reserved.
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I. INTRODUCTION Sporadic aflatoxin contamination of corn (Zea mays L) grain has probably been a common occurrence since corn was first used as human food or animal feed, Contamination is chronic in warm, humid, tropical, and subtropical corngrowing environments. These areas have favorable conditions for growth of the Aspergillusflavus group of fungi, mainly the A . flavus or Aspergillus parasiticus species that elaborate the four principal aflatoxins B , , B,, G,, and G,. Aflatoxin B , has been classified as the most potent naturally occurring carcinogen known (Squire, 1981). The toxicoses produced by moldy food and feed containing aflatoxins are now understood and the toxins and related compounds have been isolated and characterized. The extent of our knowledge of aflatoxins can be primarily attributable to a tremendous expansion of aflatoxin research during the last 20 years. The knowledge accumulated during those two decades launched us into a new era of multiple challenges (Stoloff, 1979), including a better understanding of the biology of the fungi involved, the mechanisms that trigger aflatoxin production, the biochemical pathways of aflatoxin production, and the genetic mechanisms that control these processes. Perhaps the greatest challenge, however, is that of developing a control package that combines the knowledge from each critical area of research to aid in the development of hybrids and cultural practices that reduce contamination of corn to a manageable level, if not completely eliminating the problem.
11. BACKGROUND AND IDENTIFICATION OF AFLATOXINS AS CONTAMINANTS OF CORN In retrospect, the characteristic symptoms of aflatoxicosis were reported in the literature since before the turn of the twentieth century (Mayo, 1891 ; Dalrymple, 1893). Mayo (1891) suggested that the toxicosis was a disease unique to horses since other farm animals did not seem to be affected. Reports on losses of horses (Shoenleber, 1906) were fairly common and the syndrome with its characteristic symptoms came to be known as “moldy corn poisoning” (Schwarte, 1938). Early reports implicated Aspergillus spp. fungi as possible sources of the problem, and discussions by Dr. Mayo (Schwarte, 1938) reported an association with the “green corn worm.” Taubenhaus ( 1 920) concluded that insects that invade the corn ear through holes in the husks contribute substantially to infection by molds. He did not suggest A. fravus as the causal agent that produced feeding problems in animals, probably because previous authors had indicated that Aspergillus spp. other than
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22 1
flavus were responsible for infecting corn and initiating disease symptoms (Dalrymple, 1893; Mayo, 189I ). Unique field conditions are required to enhance fungal infections of the corn ear, allowing contamination to become a serious problem during feeding and storage (Koehler, 1938, 1942). Moisture, temperature, and insects were demonstrated to be important factors influencing ear infections, but very little importance was placed on Aspergillus spp. as contributing to the disease symptoms resulting from eating infected ears. Reports of aflatoxicosis in animals by Sippel et af. (1953) and Burnside et a f . (1957) initiated research that began to focus on the Aspergilli as a primary source of the feeding problems. However, contamination was considered as being limited to stored corn (Quasem and Christensen, 1958) and the syndrome in livestock was referred to as “symptoms of disease” rather than toxicity (Burnside et al., 1957). A serious effort to identify the “causal agents” of toxicity occurred as a result of an outbreak of turkey “X” disease during 1960 in England (Blount, 1961), and identification of the aflatoxins initiated an entirely new area of scientific investigation, that of mycotoxicology. Several animal species were soon reported as being adversely affected by ingestion of aflatoxin, including cattle (Garrett et af., 1968), which were also reported by Burnside et al. (1957).
A. DETECTION AND QUANTIFICATION OF AFUTOXINS Methods for the detection and quantification of aflatoxins in peanut were developed in the early 1960s and reported by Coomes et al. ( 1964). The basic procedures for detection and quantification were modified for use on corn and other crops; the early developments were reported by Pons and co-workers (Pons et af., 1966). The procedures facilitated several surveys of marketed corn in the mid-to-late 1960s in which low amounts of aflatoxin were found (Shotwell et a l . , 1969b, 1970, 1971). Field studies in 1971 and 1972 identified aAatoxin contamination of corn as a preharvest problem (Anderson et d.,1975). Studies followed that were designed to investigate the extent of preharvest contamination in com-growing regions of the United States (Lillehoj et a l . , 1975d). Field contamination was judged to be more serious in the southern corn-growing regions than elsewhere in the United States. The difference between regions was partially attributed to ear-feeding insect activity (Widstrom et a l . , 1976). Lillehoj and co-workers laid groundwork for the investigation of genetic differences among hybrids and kernel starch types (Lillehoj et al., 1975d, 1976~). A realization that preharvest aflatoxin Contamination of corn also placed a major portion of the Corn Belt at risk led to a status report with possible solutions being presented to the 30th Annual Corn and Sorghum Research Conference in 1975 (Lillehoj and Zuber, 1975). Another status report by Zuber and Lillehoj
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( 1979) introduced control measures including genetic resistance to infection and contamination, reduction of plant stress, and insect control. The second status report was prompted in large part by the calamitous contamination of the 1977 corn crop, especially in the South (Wilson et al., 1979). The high aflatoxin contamination of the 1977 crop was a rare occurrence; however, the reduction in contamination for subsequent years (McMillian et al., 1980b, 1985b) must not be viewed as a diminution of the problem, but rather as fluctuations of chronic contamination that is always a threat in limited areas of the southern corn growing region of the United States (Table I). Indirect methods were often used for determining the presence of aflatoxins in corn fed to animals prior to the 1960s. Standardized procedures to analyze for aflatoxin content in feeds began to emerge by the middle of the next decade (Pons, 1976). Detection of aflatoxins in feed was often done by bioassaying sensitive animals such as ducklings (Sargeant ef al., 1961), chicks (Brown and Abrams, 1965), and rainbow trout (Brekke et al., 1977).
Table I Average Yearly Levels of Aflatoxin Contamination for Corn Grown in the Coastal Plain of Georgia, 1977-1994
Year I977 I978 1979 1980 1981 I982 1983 I984 1985 1986 1987 I988 I989 I990 1991 1992 I993 1994
Concentration of aflatoxin (ng g-') 622 57 68 204 91 92 I28 37 48 190 82 137 26 217 39
35 71
6
Source: McMillian et a/. (1985b). and unpublished
data.
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An association between bright greenish-yellow fluorescence (BGYF) and the presence of aflatoxin exists, but the relationship is not considered reliable since the BGYF is due to kojic acid, a secondary metabolite of A . flavus (Marsh e t a / ., 1969). The simplicity and inexpensiveness of the BGYF test attracted its use, especially before accurate measurements for aflatoxin were available. Occasional successes when using BGYF (Shotwell et al., 1972, 1975b; Rambo et al., 1975) resulted in a suggestion for its use as a presumptive test (Shotwell and Hesseltine, 198 I ) for initial elimination of some samples in surveys. Attempts to use BGYF as a routine indicator of toxin contamination have met with limited success (Shotwell et al., 1975b), however, and perhaps have been inconclusive (Dickens and Whitaker, 1981), because many isolates ofA.Javus do not produce aflatoxin but do produce BGYF (Fennel1 er d., 1973). The variability associated with both testing for aflatoxins (Whitaker et a / . , 1979) and determinations of BGYF particles (Calvert ef al., 1983) certainly contributes to their frequent failure as indicators of one another. The BGYF phenomenon has also been associated with insect damage, yet another reason why it has been ineffective as a precise indicator of aflatoxin contamination (Rambo et al., I974b; Kwolek and Shotwell. 1979). Techniques developed by Pons et al. (1966, 1973) for determining aflatoxin content in other agricultural products were gradually improved by slight modifications (Stubblefield, 1979). Variations of the minicolumn technique, introduced by Holaday (1968) were also utilized in the years following (Romer et al., 1979). By the early 1970s, Detroy et a/. (197 I ) and others (Changes in Official Methods of Analysis, 1972) had developed thin-layer chroniatographic (TLC) quantification techniques. Liquid chromatographic (LC) and TLC procedures complement one another (Trucksess and Wood, 1994) and are among the important official standard methods for analyses of aflatoxins (Official Methods of Analysis of the AOAC, 1975). The TLC method has been modified, with success, to meet the needs of laboratories in developing countries (Guzman de Pefia er a/., 1992). Good resolution of aflatoxins is also achieved by the LC method, a high-pressure liquid chromatography (HPLC) procedure of Pons ( 1976), later adapted for use on corn (Pons, 1979) and modified by Thean et a / . (1980). A comparison among blind samples at several laboratories was made by Park ef al. ( 1990) in an effort to standardize procedures. The comparison of the LC and TLC methods resulted in a conclusion that TLC tended to overestimate concentrations when amounts of aflatoxin were less than 20 ng g - I (Beaver et a l . , 1990). Other methods of aflatoxin analysis have been proposed, including the fluorometric-iodine method of Davis and Diener for regular evaluations (Davis and Diener 1979a) and for rapid screening (Davis and Diener, 1979b) and several immunochemical methods (Chu, 1990). The immunochemical methods have become popular recently because results can be obtained more quickly and at less cost than for other methods (Trucksess and Wood, 1994). The immunoaffinity
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column procedure developed by Trucksess et al. (1989, 1991) has been adopted by numerous laboratories in industry as well as in other research facilities. The method was compared with liquid chromatography for accuracy and sensitivity by Beaver et al. (1991) and found to be reliable at aflatoxin concentrations greater than 5 ng g-1.
B. THESCOPEOF CORNCONTAMINATION The first surveys conducted on grain to test for the presence of aflatoxin were made on samples obtained from large marketing centers in the United States: Chicago, Illinois, New Orleans, Louisiana, and Omaha, Nebraska. The initial survey by Shotwell et al. (1969a) did not include corn, but their finding that corn was an excellent substrate for A. flavus resulted in corn being included in subsequent surveys (Shotwell et al., 1969b, 1970, 1971). Only 48 of 2117 corn samples were found to be aflatoxin positive, and of those, only four contained more than 20 ng g-I aflatoxins. Reports of aflatoxicosis in North Carolina broiler chickens (Smith and Hamilton, 1970) and toxic hepatitis of swine and cattle in the southern United States (Wilson et al., 1967) prompted Shotwell e? al. (1973) to examine southerngrown corn from commercial markets. Aflatoxins ranging from 6 to 348 ng g-I occurred in 2 1 of the 60 samples collected. A pattern of heaviest contamination in samples from the southeastern states began to emerge. One of six samples of 1971 white corn grown in the southeastern comer of Missouri had adatoxin contamination of 400 ng g-l (Lillehoj et al., 1975a) and 8 of 163 samples grown in southern Indiana in 1972 contained aflatoxin-contaminated kernels (Rambo et al., 1974a). Documentation of preharvest contamination of corn was obtained from the 1972 crop grown in southwest Georgia (Anderson er al., 1975) when aflatoxin contamination was found on randomly sampled ears of both white and yellow corn. Sporadic reports of aflatoxin contamination in midwest corn began to appear in the mid 1970s (Rambo et al ., 1974a; Riesselmann and Doupnik, 1975; Lillehoj et al., 1976b). The reports were not localized or confined to any type of corn in that contamination of 1972 dent corn was determined in 8 of 163 samples in Indiana (Rambo et al., 1974a), as much as 30 ng g-' aflatoxins were found in 1973 Nebraska popcorn (Riesselman and Doupnik, 1975), and I1 of approximately 6000 ears of I975 Iowa corn had single-ear Contamination of 1- 1560 ng g-' (Lillehoj et al., 1977). When corn from several locations was compared, however, the southern locations, especially Georgia, North Carolina, and Texas, always showed the heaviest contamination (Lillehoj et al., 197%; Zuber et al., 1976). Samples of 1973 and 1974 corn grown in South Carolina had 24-209 and 0-281 ng g-I aflatoxins, respectively, while 1974 Florida corn was contami-
THE AFLATOXIN PROBLEM WITH CORN GRAIN
225
nated with 3-1218 ng g-1 (Shotwell et al., 1977; Lillehoj et ul., 1976~).One instance of contamination found in white corn grown in 1971 in southeast Missouri drew much attention because white corn is often used for food products. Among 1283 truckloads, 165 had greater than 20 ng g-' and 29 loads exceeded 100 ng g - l aflatoxin (Shotwell et al., 1975a). National attention was focused on the aflatoxin problem in 1977 when heavy contamination occurred throughout the southeastern states. Formal reports on the extent of contamination of the 1977 crop were published for the states of Alabama (Gray et al., 1982), Georgia (McMillian et al., 1978), and North Carolina (Hesseltine et al., 1981), although corn grown in the other southern states was similarly affected. The 1977 midwestern crop, however, had a very limited amount of contamination (Shotwell et al., 1980), and five of eight states in the midwest were reported to have escaped completely, based on the survey. The state of Georgia continued to monitor its corn to determine preharvest contamination, and later reported survey results of 1977 through 1982 (McMillian et al., 1985b). This survey clearly demonstrated that the preharvest problem in Georgia is chronic, but contamination vanes greatly from year to year (Widstrom et a l . , 1984b); that is, contamination was heavy in 1977 and 1980, moderate in 1981 and 1982, and significantly less in 1978 and 1979 (Table I). Preharvest aflatoxin contamination of the corn crop is chronic only in the South and Southeast, and is not considered a threat in many areas of the United States, but it has been reported in no fewer than 21 states (Wilson and Payne, 1994). Contamination of the corn crop in other countries is quite possibly more serious than that in the United States because a greater proportion of the corn grown in many countries is used as a human food source (Jelinek, 1987; Jelinek et al., 1989). Campbell and Stoloff (1974) discussed the implications of mycotoxins on human health while others studied their effects on domestic animals (Wilson et af., 1967) and other test animals such as the rat (Carnaghan, 1967). The concern, therefore, is worldwide for two reasons: first, because corn is shipped routinely to importing countries, and second, because corn produced and consumed within numerous countries throughout the world has been reported as contaminated (Wood, 1989). The Philippines (Ilag et al., 1976), France (Galtier et ul., 1977), Yugoslavia (Balzer et al., 1977; Durakovic ef al., 19841, Egypt (Farag et al., 1980; Qutet et al., 1983), Mexico (Martinez, 1979; Zuber et a l . , 1986), Brazil (Sabin0 et al., 1989, Zuber et a l . , 1986), Haiti (Castor et al., 1987), India (Bilgrami et al., 1981a; Zuber et al., 1986), Australia (Blaney, 1981), Thailand (Shank et al., 1972a; Zuber et al., 1986), and Bolivia, Colombia, and Costa Rica (Zuber et al., 1986) have all reported contamination of their corn. Africa, however, probably has a more widespread problem than any other continent and has reported contamination in seven of its countries: Uganda (Alpert el al., 1971), Zambia (Lovelace and Nyathi, 1977), Kenya (Ngindu et ul., 1982), Nigeria (Okoye, 1986), Mozambique (Purchase and Goncalves,
226
N. W. WIDSTROM
1971), Tanzania (Seenappa and Nyagahunga, 1982), and Ghana (Zuber et al., 1986). The greatest concern is for people in countries that grow much of their own corn, use corn as a staple food in their diet, and have no organizational setup to monitor corn grown and consumed at the local level. These areas include Mexico and several countries in Africa and Central and South America.
111. CONDITIONS IMPACTING ASPERGILLUS F L A W S GROUP INFECTION AND AFMTOXIN ACCUMULATION Numerous environmental and cultural conditions have an influence on infection and aflatoxin production processes which occur in the corn kernel when exposed to spores of the A . f l a w s group. The gamut of these conditions has changed little since being reviewed by Lillehoj (1983), but our understanding of the importance of each and how they interact has received considerable attention in research conducted since (Widstrom, 1992). A thorough knowledge of conditions that we cannot control, which exist prior to planting a corn crop, as well as those over which we have some control may indeed make the difference between whether a crop escapes or suffers contamination by aflatoxin.
A. THEINFLUENCE OF FACTORS BEYOND GROWER CONTROL Assuming that a grower is restricted to growing corn in a limited area, weather and soil type factors are quite firmly fixed. The grower, therefore, must learn to cope with or compensate for these uncontrollable conditions, if they are unfavorable to the production of a contaminant-free crop.
1. Weather The prevalence of Aspergillus ear molds has long been associated with dry weather, although wet weather provides no assurance that ear molds will not be present (Taubenhaus, 1920). His findings are in agreement with the fact that the Aspergillus as a group of fungi seem to appear before other kinds of fungi when kernels have a low moisture content, 5140 g H,O kg-' dry matter (Koehler, 1938). The first confirmation of the incidence of aflatoxin being found in preharvest grain also pointed to a higher incidence of aflatoxin contamination of corn grown in the warmer regions of the United States (Anderson et al., 1975). The report by Anderson et al. (1975) initiated a number of studies to more clearly define weather-related differences in aflatoxin incidence and amounts as they relate to varied corn-growing regions in the United States (Lillehoj et al.,
THE AFLATOXIN PROBLEM WITH C O W GRAIN
227
1975d; Zuber et ul., 1976). The general conclusions of these studies were that both incidence and amounts of aflatoxin in the corn sampled increased from north to south and that the increase in southerly locations in the United States was definitely related to temperature and possibly also related to regional differences in precipitation (Lillehoj et ul., I978b). Weather-related regional differences within states were suggested in many of the earlier reports as a reason why differences in infection and incidence of BGYF and aflatoxin can occur between regions (Lillehoj and Hesseltine, 1977). Several surveys were conducted on corn from the Midwest during the 1960s and early 1970s (Shotwell, 1977). Most of the surveys located a few samples with low levels of aflatoxin, with the exception of the white corn harvest of 1971 in southeastern Missouri. Nearly one-third of the truckloads sampled from stored corn from this harvest had detectable levels of aflatoxin. Samples of southerngrown corn, however, often were found to have contamination incidences of 4050% (Shotwell, 1977). Apparent regionalization of the heaviest contamination encouraged recommendations by the extension service that growers should make a serious effort to avoid drought stress during production of the crop (Duncan, 1979). Any notion that aflatoxin contamination was nearly always confined to the southern corn-growing region was dispelled by results from surveys of the 1977 crop when more than 18% of 87 samples from the drought-stressed crop in central Iowa had amounts >20 ng g-1 (Zuber and Lillehoj, 1979). While heavy contamination occurred locally in 1972 in the Midwest, levels of aflatoxin in 1977 southern-grown corn can be described as no less than disastrous (Wilson et a/., 1979; Manwiller and Fortnum, 1979; Gray ef a/., 1982). From that point forward, research on a solution or control of the problem was begun in earnest. The problem was not so serious in 1978 (McMillian et al., 1980b), but 2 successive years of heavy contamination in I980 and 198I Georgia corn (McMillian ef al., 1985b) convinced any remaining skeptics that chronic contamination, at some level, existed for southern-grown corn. One common denominator of field studies has been that high temperatures are associated with greater amounts of aflatoxin contamination of field samples (Jones et al., 1980; Zuber et al., 1983; Hill et a / . , 1985). High temperatures are also nearly always an important component of drought and the plant stress associated with drought. Drought stress has been commonly given as a major component of contamination in those years when aflatoxin levels were high (Davis et a / . , 1985). The persistence of conclusions that include drought and plant stress as major components of contamination is not surprising, since detailed studies of weather-associated factors have concluded that high temperature and low humidity, that is, evaporation or transpiration losses, are significantly correlated with heavy contamination of corn sampled directly from the field at harvest (Widstrom et a/. , 1990). Weather variables, in addition to temperature
228
N. W. WIDSTROM
and moisture, have been suggested as having an influence on aflatoxin production (Fortnum, 1987) and a study by McMillian et al. (1985) illustrated the complex interrelationships that exist among moisture, insect damage, temperature, and plant resistance to aflatoxin production. It is, therefore, not surprising that some investigations do not always indict drought stress and temperature as being the dominant factors producing aflatoxin contamination (Stoloff and Lillehoj, 1981). High temperatures are consistently found to be an important factor influencing aflatoxin contamination and fungal growth when tests are conducted under controlled conditions (Thompson et al., 1980; Holmquist et al., 1983; Hill et al., 1985; Kingsland, 1986; Wieman et af.,1986; Payne et af., 1988b). Modifications of temperature and plant stress through irrigation or other means are always important components of recommended control measures, even when genetics and host plant resistance are principal emphases (Zuber and Lillehoj, 1987; Widstrom, 1987). The first attempts at correlating specific weather within a weekly or monthly time frame with the incidence of aflatoxins were made by Sisson (1987). The importance of temperature and humidity within time windows was corroborated by Widstrom et al. (1990), who suggested that, contrary to the standard recommendations for early planting, those plantings in the deep South are believed to be at higher risk for aflatoxin contamination than those which are delayed to change the time period for grain-filling. The most recent comprehensive review (Payne, 1992) states that no control strategy is completely effective for presently grown commercial hybrids when environmental conditions are extremely favorable for growth of the fungus.
2. Edaphic Factors Recommendations made by the extension service to help minimize aflatoxin contamination of the corn crop have always included adequate fertilization and irrigation to provide a root zone in the soil that will impose a minimum of stress on the plant (Georgia Extension Aflatoxin Committee, 1978; Glover and Krenzer, 1980). The suggestion of altering edaphic factors to reduce aflatoxin contamination of the crop, whether through fertilization, irrigation, or cultivation, may relate to the fact that the soil serves as a repository for the spore load imposed on the crop to be planted. Cultivated soils seem to carry higher spore concentrations than others (Angle, 1987), probably increasing the likelihood of exposure to infection and necessitating production of healthy plants that will resist infection. Angle (1987) also demonstrated that degradation of aflatoxin occurred more slowly in silty clay loams than in fertile silt loam soils; however, no information was given on inactivation of A . flavus spores in those soils. In one of the first studies involving geographical differences, contamination was primarily attributed to the differences in weather and plant factors, but soil factors were probably also involved (Lillehoj et ul., 1975d). The interrelation-
THE AFLATOXIN PROBLEM WITH CORN G M N
229
ships among factors have not been fully sorted out in that the choice of fields may, for example, be based on whether soils are sandy (droughty) since they are more likely to produce stress on corn than on sorghum (Jones, 1987). The importance of the soil as a source of inoculum has been well documented (Lillehoj et a / . , 198Od) and has been included in reviews when control measures were discussed (Widstrom et al., 1984b; Zuber and Lillehoj, 1987; Wilson et al., 1989b). Martyniuk and Wagner (1978) demonstrated that management systems such as continuous cropping have an impact on the quantity and quality of microflora. Some tillage studies have provided mixed results; subsoiling in North Carolina reduced aflatoxin contamination (Payne ef al., 1986), but differences due to one, two, or three cultivations for weed control were nonsignificant in India (Bilgrami et a / . , 1992). Experiments involving fertilization and irrigation effects have been more definitive than those on tillage, and recommendations for cultural control of aflatoxin contamination of corn always include the need for maintaining adequate fertility and moisture in the soil profile (Duncan, 1979; Smith, 1981; Jones, 1987; McMillian et a l . , 1991). Good nutrition of corn reduced contamination when stresses by other factors were not present (Wilson et al., 1989a). Similar conclusions were drawn from two independent studies in North Carolina (Jones and Duncan, 1981; Payne et al., 1989). The interrelationship between good fertility, available soil moisture, and other factors has been the subject of several studies, and interaction among the influencing factors should be expected (Jones et ai., 1981; Smith and Riley, 1992). Tremendous differences from year to year that were encountered in some of these studies (Fortnum and Manwiller, 1985) were undoubtedly responsible for conflicting views as to their importance to contamination of corn and to recommendations given for management. For example, Jones et al. (1981) found the least contamination in early plantings while samples from early plantings grown by Smith and Riley (1992) had significantly larger amounts of aflatoxin than late plantings. Recommendations for adequate irrigation are, of course, standard for grain production and are merely reinforced as far as prevention of aflatoxin contamination is concerned.
B. PREPLANTING CONSIDERATIONS An adequate knowledge of the crop and management history of the area on which corn is to be grown is a necessary prerequisite to improving the probability for producing an aflatoxin-free grain crop. The adage that “an ounce of prevention is worth a pound of cure” is definitely applicable to corn grown for grain. Decisions regarding where, when, and what to plant can make the difference between success and failure in producing profitable crops free of contamination. Preplanting decisions are impossible to change after the crop has emerged; there-
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N. W. WIDSTKOM
fore, careful preseason planning is critical to reducing contamination of a corn crop.
1. Soil Testing Soil testing is critically necessary for evaluation of fertilizer needs of areas where corn is to be grown. Corn requires a higher soil pH and more nitrogen than most other crops. Maintaining an adequate soil pH and nitrogen supply is difficult in the South because the area in general has greater rainfall than the Corn Belt, and sandier soils, lowering the pH and leaching soluble nutrients from the root zone (Aldrich et a l . , 1975). Adjustments of soil pH with lime application should be made well in advance of planting, before the pH reaches critically low levels, because adjustment of pH due to liming is usually not effective during the same growing season it is applied, and the crop will have matured before breakdown of the lime can have any noticeable effect on the pH. Soil testing and subsequent preplant application of lime and fertilizer to alleviate deficiencies in pH, nitrogen, and other plant nutrients enable the grower to get his crop started with a minimum of the stresses that have been reported to predispose the crop to aflatoxin contamination (Jones and Duncan, 1981). Deficiencies of other major (phosphorus and potassium) and minor elements, as determined by soil test, can normally be remedied at planting by application of a complete fertilizer blended by commercial dealers especially for corn production.
2. Crop Rotation Quantitative differences in soil microflora, including A , flavus, have been found in soils that have been placed under continuous cropping of corn (Martyniuk and Wagner, 1979). Conventional tillage practices in a red clover-wheat rotation yielded soil samples with 256 propagules of A. flaws and A . parasiticus per gram of soil. Subsoiling, which is a more frequent practice in some rotations than in others, has been demonstrated to be beneficial in reducing aflatoxin contamination of corn (Payne et al., 1986). Specific crop rotations have not been compared sufficiently to warrant recommendations other than to encourage basic rotation principles, such as the avoidance of continuous cropping. The use of cultural practices and rotations that optimize production are those that minimize contamination and infection (Widstrom, 1992), and are routinely recommended, having proven to be effective in practice (Wilson e t a ! . , 1989b). Another aspect of the soil microflora associated with rotation, continuous cropping in particular, is the formation of sclerotia on crop debris after harvest (Wicklow et al., 1982, 1984). The resistance of these fungal structures to decomposition and quick germination could prove them to be an abundant source of inoculum in the following year. Sclerotia from certain A. flavus isolates have
THE AFLATOXIN PROBLEM WITH CORN GRAIN
23 1
been shown to germinate in the field, providing inoculum prior to silking of the corn crop (Wicklow and Wilson, 1986).
3. Hybrid Selection The selection of an appropriate hybrid for planting is a vital part of the grower’s management program to minimize aflatoxin contamination of his corn grain crop. An earnest search was begun for hybrids that would restrict or prevent the infection by A . jlavus and accumulation of aflatoxin in their grain (LaPrade and Manwiller, 1976, 1977; Lillehoj et a / . , 1976~;Widstrom et al., 1978), as soon as a preharvest contamination problem was documented (Anderson et a l . , 1975). Although most investigators agreed that differences existed among hybrids for resistance to aflatoxin formation, questions regarding the source or causes of that resistance were still being extensively discussed. Several of the early studies that included numerous hybrids were designed to answer questions other than whether hybrid differences were worthy of pursuit as a solution to the problem (Lillehoj et al., 1982b, 1983a); therefore there were difficulties in acquiring definitive information on hybrid differences (Lillehoj and Zuber, 1981). Some reports involving large numbers of hybrids suggested that no differences for resistance existed among hybrids (Davis et al., 1985). Prior to 1980, most recommendations regarding hybrid selection were guarded in that growers were encouraged to use adapted hybrids (Zuber, 1977; Zuber and Lillehoj, 1979) alluding to the avoidance of stress during development. Early reports associated field infection of corn ears by Aspergillus spp. with insect infestations (Taubenhaus, 1920). The report documenting preharvest contamination of corn by aflatoxin also associated the problem with insect damage (Anderson et a l . , 1975). Several lepidopteran insects were identified as contributors to contamination, but the European corn borer (Ostrinia nubilalis, Hiibner) was associated with the highest concentration of aflatoxin (Widstrom et a l . , 1975). Confirmation of the finding by Fennel1 et al. (1978) pointed toward insect resistance as a critical factor in the choice of hybrids by the grower. Maize weevils (Sitophilis zeamais Motschulsky) have also been shown to effectively transport the fungus into corn ears. Therefore, resistance to weevils as well as other insects must be a consideration in hybrid selection (McMillian et a l . , I980a). Husk tightness has always been a component of field resistance to insects and as such has been shown to be an important contributor to reduced aflatoxin contamination (Barry et a / . , 1986; McMillian et a l . , 1987; Widstrom et a l . , 1994). While Corn Belt hybrids have loose husks for quick dry-down, most southern-bred (adapted) hybrids in the southeastern United States have sufficient husk coverage and tightness to discourage insect invasion. Husk traits are consequently very important in the hybrid selection process. Genetic differences for both hybrids and inbreds in A. j a w s infection or
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N. W. WIDSTROM
aflatoxin accumulation that are independent of resistance to insects and husk traits have been clearly demonstrated (Zuber et al., 1978; Widstrom et al., 1987; Scott and Zummo, 1988). Recent recommendations specifically suggest that emphasis must also be placed on genetic differences that have as their basis something in addition to adaptation, husk traits, and resistance to ear feeding insects (Widstrom and Zuber, 1983; Zuber and Lillehoj, 1987; McMillian et al., 1991). The grower, at this point, is still faced with basing his selection of a hybrid on adaptation, husk traits, and resistance to insects, since hybrids with chemically based or other resistance mechanisms have not yet been developed. a. Plant Population Recommendations to minimize risk of aflatoxin contamination of corn routinely caution against planting populations in excess of those optimum for yield of a given hybrid under expected growing conditions (Georgia Extension Aflatoxin Committee, 1978; Duncan, 1979). The recommendations are made based on avoidance of stress that is known to influence the infection and contamination processes (Wilson et al., 1989b). There is a dearth of information on the influence of plant populations and opposing hypotheses have been proposed for possible effects of population density (Jones, 1987). High plant populations encourage stress due to competition for nutrients and water, while on the other hand they also produce heavy canopies that may reduce exposure of maize ears and silks to airborne spore load. An additional factor associated with high plant populations, introduced by McMillian et al. (1985a), related to the amount of free water and period of time that free water is present on ears and silks during kernel maturation. Ears that were repeatedly exposed to free water over a 4-week period to simulate heavy dews common in the Southeast sustained elevated levels of aflatoxin contamination. The finding is consistent with other recent studies that have correlated net evaporation rates during kernel maturation with the amount of aflatoxin found in those kernels (McMillian et a l . , 1985b; Widstrom et al., 1990). The lowest plant population sustained the highest level of aflatoxin contamination in a crop grown during monsoon when rainfall amounts varied from 5 to 10 cm week-' (Bilgrami era/., 1992). Temperature rather than moisture or humidity was probably the determining factor for toxin contamination during this study in which varieties rather than hybrids were compared. Optimum populations for hybrids under varied conditions are often listed on the bags of commercially produced seed corn. Exceeding those recommendations, especially for noninigated corn, can be an invitation to A . Javus infection and aflatoxin contamination. Lower plant populations should be seriously considered also when the crop is planted in soils with high sand content since these soils have lower waterholding capacity and are more likely to impose drought stress than soils with high clay content.
THE AF'LATOXIN PROBLEM WITH CORN GRAIN
233
b. Planting Date The earliest studies conducted in 1977 in Georgia (Georgia Extension Aflatoxin Committee, 1978) and in 1978 and 1979 in North Carolina (Jones and Duncan, 1981; Jones et a f . , 1981) indicated that the latest planting dates produced the highest amounts of aflatoxin contamination in grain samples. Heavier contamination of South Georgia plantings in 1974-1976 than those further north was partially attributable to later planting dates, but the differences were confounded with location effects (Widstrom et a / ., 1978). Recommendations were made for planting early to minimize contamination, based on these early experiments, even though planting date was not the primary factor for investigation in the studies (Duncan, 1979; Glover and Krenzer, 1980). Mixed signals regarding the relationship between planting date and aflatoxin contamination (Widstrom et a / . , 1978; Lillehoj et a/., 1980b) suggested that planting date means were being influenced by other factors and possibly may be location specific. Consequently, specific planting recommendations were avoided (McMillian et al., 1985b) until a more complete pattern emerged from the then unpublished work in Georgia (Widstrom et a / ., 1990). Recommendations to delay planting as long as possible because later plantings on the Coastal Plain were subject to lower contamination levels began to appear as early as 1984 (Widstrom er a / . , 1984b) and have been continued since (Wilson et al., 1989b; McMillian et al., 1991; Widstrom, 1992)). Since delayed plantings produce lower grain yields, a grower must weigh contamination risk against lower yields when making a planting date decision. Additionally, the effect of reduced contamination for later plantings may not hold for other locations where temperature, soil type, etc., are different. Studies in Louisiana, however, showed a lower level of aflatoxin contamination for the latest of three planting dates than in either of two plantings 10 and 20 days earlier (Smith and Riley, 1992). The basic tenet for late planting is to time planting so that the critical grain filling period, beginning at 20 days after silking, occurs after the highest seasonal temperatures and period of net evaporation (Widstrom et al., 1990). The data showing the trends in sample contamination of wound-inoculated ears for planting dates obtained over several years are given in Table 11. A general reduction in aflatoxin Contamination of 150-200 ng g-I occurs from early to late for each approximate 15-day planting interval. c. Resistance to Insects and Diseases Insect damage recorded in aflatoxin studies revealed that the highest levels of aflatoxin contamination were usually associated with heavy insect damage (Lillehoj et a / . , 1 9 8 0~;Wilson et al., 1981a). Lillehoj and Hesseltine (1977) suggested that insects were important as carriers of aflatoxin-producing fungi. Specific insects had been identified as being associated with the presence of A. ,flovus, i.e., corn earworm (Lillehoj et a / . , 1976d) and European corn borer
N. W. WIDSTROM
2 34
Table I1 Planting Dates, Grain-Fdling Periods, and Geometric Means for Aflatoxin Concentrations of Wound-InoculatedCorn Samples at TiPton, Georgia, 1982-1987"
Planting date
Grain filling period
27 February 18 March 31 March 15 April 1 May 15 May 29 May 13 June 30 June 14 July 29 July
I3 June- I9 July 21 June-28 July 29 June-8 August 8 July- 15 August 19 July-24 August 2 August-7 September 13 August- I 8 September 28 August-3 October I3 September- 18 October 29 September-5 November 12 October- 17 November
~~
Average anatoxins (ng g - I P I939 2018 1604 898 412 909 104 171
459 67 33
~
Adapted from Widstrom (1992). Geometric means are the antilogarithms of the logarithmic means for aflatoxin concentrations.
(Lillehoj et a l . , 1976b). A comparison of several insect species and their impact on aflatoxin contamination made by Widstrom et al. (1975) revealed that the European corn borer was among the most effective in exacerbation of the infection and contamination processes. This insect may be an important factor in preharvest contamination during sporadic outbreaks of contamination that occur in the Corn Belt (Guthrie et a l . , 1982). The maize weevil, a late-season, preharvest pest of corn in the Southeast, has been identified as an effective vector of A. j h v u s , while the wheat curl mite (Eriophyes tulipae, Keifer) was determined to be ineffective as a vector (Barry et al., 1985). Nitidulid beetles (Nitidulidae: Coleoptera) were associated with A . f l a w s infection of wounded kernels (Lussenhop and Wicklow, 1990), and A. flavus contamination of corn earworm moths was significantly correlated with aflatoxin contamination of grain sampled from the Coastal Plain of Georgia over a 6-year period (McMillian et a l . , 1990). No hybrid is resistant to attack by all of the insects that have been implicated as contributors to aflatoxin contamination, but most states have performance bulletins that identify the hybrids most resistant to insects in their test areas to assist growers in hybrid selection. Husk tightness is an important component of plant resistance to most earfeeding insects, and several studies were initiated to determine if that trait was also critical in reducing aflatoxin Contamination. A near-linear relationship be-
THE MLATOXIN PROBLEM WITH CORN GRAIN
235
tween husk tightness and grain contamination was found among five hybrids varying in husk tightness and evaluated under inoculated conditions when infected by corn earworm and European corn borer (Barry et ul., 1986). Similar results were obtained by McMillian et al. (1987) and Widstrom et al. (1994), reaffirming the importance of husk cover in regions where insects and/or contamination are chronic problems. The experiments with hybrids varying in husk tightness provided an alternative explanation for experiments that had utilized hybrids from varied locations. Differences in contamination that initially had been ascribed to differences in adaptation could be more logically ascribed to variation in husk cover. In fact, most of the differences in contamination were probably due to husk cover since southern hybrids have much better husk protection than those developed for the Corn Belt. Lillehoj ei al. (1976~)determined that hybrids adapted to the South and grown in the South had lesser amounts of aflatoxin contamination than Corn Belt hybrids grown at the same location. Resistance to insects that restricts insect activity in the ear, whether due to husk traits or other inherited resistance traits, will reduce the resulting amount of aflatoxin contamination in the grain. This is especially true for European corn borer, which is also a notorious leaf-feeder (Lillehoj et a/., 1982a). Ear damage by insects, when present, is nearly certain to be associated with the level of aflatoxin contamination found in the grain, and has probably been an influencing factor determining hybrid differences for contamination as a function of husk coverage and plant resistance to those insects (Zuber er a / ., 1983; Barry et al., 1992). Selection of a hybrid with resistance to diseases is obvious as a means to maximize yields, but it is a choice that will also reduce the risk of aflatoxin contamination. The predisposition of seed to aflatoxin contamination by ear rots, such a5 Helrninthosporium maydis, was first noted in grain samples from Georgia in 1970 and 1971 (Doupnik, 1972). Aspergi/lus$uvus is often considered to be an ear-rot organism in its own right, and as such is inextricably confounded with damage by the ear-rot complex infecting corn ears in the field. Some investigators have chosen to evaluate A . flavus infection in that context (Campbell et al., 1993). Inoculation and evaluation techniques similar to those used for other ear rots have been employed for A . flavus (Campbell and White, 1994), but it must be remembered that aflatoxin concentration is the ultimate trait of interest. Earrot evaluation procedures have been successful in identifying some hybrids with resistance, even though many evaluations have been made only on a visual basis. Attempts to separate the influence of organisms other than A. Javus that are present on the ear have been generally confined to viewing them as competitors for substrate and as such have been considered as a possible means of control. This concept will be discussed in a subsequent section of this chapter. With respect to grower selection of a hybrid, most separations of resistant types have been based on aflatoxin contamination at harvest (Widstrom et al., 1978; Damah
236
N. W. WIDSTROM
et al., 1987; Kang et al., 1990; Wallin et al., 1991) or the percentage of infected kernels at harvest (Tucker et al., 1986; Scott and Zummo, 1990a). Both resis-
tance evaluation methods have been validated as identifying many of the same resistant germplasm sources, and therefore results from studies designed to answer other questions seem compatible, independent of the evaluation method used (Widstrom et al., 1978, 1984c; Scott and Zummo, 1990a, 1994; Scott et al., 1991). The greatest difficulty at this time is availability of hybrids that have moderate levels of resistance to insects, diseases, and aflatoxin contamination, not in identifying those hybrids.
IV. MANAGING CONDITIONS DURING PLANT GROWTH AND EAR DEVELOPMENT The ultimate management tool for agronomic problems, including injury or product quality, in plants is that of manipulation through plant breeding or other genetic techniques designed to give control during plant growth and development. Initial efforts to locate and utilize genetic resistance to aflatoxin contamination in the corn plant were not very successful in that experimental results were nonconclusive (Zuber, 1977; Widstrom et al., 1978; Lillehoj et al., 1980a). Several management practices have been proposed to supplement the genetic sources of resistance in the plant, although most investigators concede that plant resistance will be a major component of any control package to reduce aflatoxin contamination (Widstrom, 1992).
A. MINIMIZING STRESSES ON THE CORNPLANT Components of any management package, whether or not perceived as temporary, include insect and disease control, and alleviation of conditions that contribute to stress on the plant (Lillehoj and Hesseltine, 1977; Zuber and Lillehoj, 1979, 1987). Stress relief has been viewed as a major component of resistance to the contamination process (Lillehoj, 1983; Fortnum, 1987). The stress component must continue to receive major emphasis (Georgia Extension Aflatoxin Committee, 1978; Duncan, 1979; Smith, 1981) throughout the life of the plant (Fig. I).
1. Irrigation One of the easiest methods of preventing stress on the plant is to avoid drought. This can be accomplished by an adequate irrigation system and is a
+
Environmental stTesses on plants, diseases, and insects
-L
Wind, water, and insect borne spores from colonized debris, sclerotia, and soil sources
,+ 20
Whorl to silking L -Infection Vegetativeperiod
I
and Colonization period-
60
L
I
t H a r v e s t period 4 4 - S t o r a g e and -b 135 utilization 120
b F t susceptible 65 time of infection 85
tl
Approximate no. of days post-planting (-Increasing
b -
damage to grain by insects
*
Increasing post-infection aflatoxin accumulation
b
figure 1 The chronology of corn kernel infection by Aspergi/lusJ?avrts and subsequent atlatoxin contamination. Source: Widstrom ( 1992).
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N. W. WIDSTROM
standard recommendation for corn grown in the South and other warmtemperature locations, especially those that have sandy soil with low waterholding capacity (N. C. Aflatoxin Committee, 1977; Glover and Krenzer, 1980, Smith, 1981). The risk for aflatoxin contamination of corn, however, seems always greatest under drought conditions, regardless of soil type (Tuite et af., 1984), and the recommendation made to avoid contamination is to alleviate stress by irrigation during the reproductive period (Jones, 1983) or adjust the planting date to move the critical period of grain filling to a period of minimum stress (Widstrom ef al., 1990). All irrigated corn production systems require some form of soil moisture monitoring to determine when irrigation is needed (Lee, 1994). Providing water to the crop in efficient amounts at the optimum time will often determine the profit margin for production; therefore, good judgment and experience are required for wise decisions regarding the best time to irrigate. Most growers, experienced or not, will rely on mechanical or electronic devices to determine when such soil moisture levels are critically low and irrigation is needed. Tensiometers or similar devices give the most reliable soil moisture measurements and can provide moisture availability at several soil depths, giving the grower adequate information to make a good decision on when to irrigate. This information, along with up-to-date weather forecasts, will maximize water use efficienCY.
The soil water tension in centibars required to call for irrigation will vary with plant stage, soil type, and the adequacy of the irrigation system (Lee, 1994). Young plants can survive slightly lower levels of moisture in the soil before irrigation is applied, while large amounts of water are needed at the critical flowering and grain-filling stages. Water demands are so high during the critical stages that some plant stress is seldom avoided, especially if temperatures are high and rainfall is limited during these periods. In general, irrigation is called for when 20- to 25-cm-depth tensiometer readings are at 20 centibars or greater. Sandy loam soils usually require 25-40 mm of irrigation when the critical soil moisture tension is reached. Heavier soils can handle slightly more and sandy soils slightly less because of a lower waterholding capacity for sandy soils. Moisture deficit is among the easiest plant stress-inducing factors to adjust and probably the most important because it significantly impacts other stresses, such as insect damage and disease expression. Sandy soils, subject to frequent moisture deficit, along with high night temperatures and greater disease and insect pressure, are the principal reasons why aflatoxin contamination of corn is chronic in the southern and southeastern United States. Numerous studies have investigated the influence of irrigation on aflatoxin contamination (Fortnum and Manwiller, 1985; Payne er af., 1986; Jones, 1987; McMillian et af., 1991; Smith and Riley, 1992). These studies, without excep-
THE AFLATOXIN PROBLEM WITH CORN GRAIN
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tion, demonstrated a net beneficial effect when irrigation was available. The benefit of irrigation cannot always be realized, however, because it is often not practical for the grower. In fact, corn is most often produced without irrigation in high-risk areas, since more than one-half of the corn acreage in the Southeast is grown under nonirrigated conditions. Alternative control measures must therefore be made available to growers for whom irrigation is either impractical or impossible.
2. Fertilization and Plant Nutrition Initial observations of an increased incidence of aflatoxin contamination in preharvest corn grown under low fertility conditions were made by Anderson et al. (1975). This study in Georgia and others have led to a general consensus that nitrogen fertilization of corn will influence aflatoxin contamination of the crop (McMillian et al., 1991), even as it influences most other plant traits. The sandy coastal plain soils of the southeastern United States are naturally very low in the highly soluble nitrogen that is critically needed for corn, a heavy user of this element (Gurley, 1965). Since a recommendation of adequate fertility is critical for obtaining good yields, no serious changes in the fertilization recommended for corn production were necessary with regard to aflatoxin contamination (Georgia Extension Aflatoxin Committee, 1978). A word of caution resulted from experiments by Wilson et al. (1989a) when they demonstrated that overfertilization with nitrogen can also increase the incidence of contamination. This effect can again probably be attributed to increased stresses on the plant and is a concern only for those who are attempting to obtain maximum yields by applying high levels of nitrogen fertilizer. Other fertilization studies have given similar results regarding the need for a supply of adequate nitrogen for the corn plant (Glover and Krenzer, 1980). No single experiment can be cited as conclusive proof of the influence of nitrogen on Contamination, since many studies also include the testing of other confounding factors (Jones, 1983; Jones and Duncan, 1981). Stresses induced by inadequate nitrogen for good plant growth are clearly a significant contributor to the contamination process (Payne et al., 1989). Lillehoj (1983) reasoned that since stress is so convincingly implicated, and inadequate fertilization does induce stress, we must include fertilization in the aflatoxin contamination equation. The nutritional status of the plant, other than that expressed by obvious deficiency symptoms and lack of vigor, has not been demonstrated to be closely associated with contamination by aflatoxin. Most nutritional factors have a high impact potential on yield and are normally addressed because of their close relationship to capacity for production. Many nutritional problems occur because of nutrient solubility that is related to pH of the soil solution. Adjustments in pH are made by the application of lime, as previously described.
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N. W. WIDSTROM
a. Deficiency Symptoms Nutritional deficiencies can usually be avoided if the appropriate fertilizers are applied in a timely manner based on soil tests. Weather or unusual edaphic conditions may induce deficiency symptoms in the corn plant due to lack or unavailability of essential nutrients. Unlike the symptoms of disease development and insect activity, symptoms of nutritional deficiency can often be remedied and the plant restored to a healthy condition, if soil pH is in an acceptable range for corn growth and weather is not extreme. Frequent field inspections (as often as twice weekly) will assist greatly in identifying plant stress due to nutritional inadequacies. Books and pamphlets are available which not only describe deficiency symptoms, but also give excellent pictorial examples to assist in diagnosis (Aldrich et al., 1975). County agents are familiar with these aids and are available to assist the grower with both the diagnosis and the remedy. Prompt attention to deficiencies will increase production and avoid the plant stress which can predispose kernels to A. flavus infection and aflatoxin contamination. b. Tissue Sampling The plant is already suffering from stress if one waits until deficiency symptoms appear. Whole plant or leaf analyses can be useful for anticipating nutritional problems if a systematic program of testing is used (Lee, 1994). This procedure is very useful after the whorl stage for systems where fertilizer can be applied through the irrigation system. A standard range of acceptable values has been established for the major elements and most minor elements at the various stages of plant growth (Smith, 1990). When samples are taken on a regular schedule, the nutritional needs of the plant can be accurately anticipated prior to stress due to nutrient deficiency. Stresses on the plant, especially during the critical grain-filling stage, are known to increase the risk of aflatoxin contamination, and a regimented system of tissue sampling can eliminate nutritional stresses during that critical period.
3. Cultivation and Weed Control Cultivation and weed control are sometimes thought of as being synonymous, but for purposes of this discussion, cultivation includes all types of tillage and/or disturbance of the soil. It seems that cultivation practices associated with any crop tend to increase the incidence of Aspergillus spp. propagules in the soil. When compared to virgin, undisturbed prairie soils which produced 0 propagules, soils under conventional tillage and a legume-grass rotation yielded 256 propaguleslgram of soil (Angle, 1987). The degradation of aflatoxin also varied from one soil to another, in that a fertile silt loam soil was more efficient than a silty clay loam soil at decomposing aflatoxin B , .
THE AFLATOXIN PROBLEM WITH CORN GRAIN
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Cultivation practices used under different rotation systems have not been shown to influence aflatoxin contamination of the corn crop (Smith, 1981), nor have the practices of conventional till versus no-till. Presumably, all tillage systems provide an adequate supply of inoculum for infection and aflatoxin contamination when environmental conditions are favorable. One tillage practice that has proven effective in reducing contamination is that of subsoiling. Subsoiling allows deeper root penetration and renders the plant less susceptible to stress under drought conditions. The apparent benefit of subsoiling is accomplished by buffering the plant against water stress (Payne el al., 1986). Subsoiling is apparently the only tillage practice proven beneficial in reducing aflatoxin contamination, although recommendations usually only refer to tillage as an influencing factor (Jones, 1987). A good program of weed control is a necessity for every successful corn growing operation. Eliminating weeds will obviously reduce water usage and assist in preventing water stress on the crop, reducing yield losses for dryland corn. As a secondary effect, good weed control will also reduce contamination by aflatoxin, and consequently recommendations for control of aflatoxin usually include judicious control of weeds by chemical or other means (N. C. Aflatoxin Committee, 1977; Glover and Krenzer, 1980). The importance of addressing weed competition with the crop in an aflatoxin control program has not been documented by any formal studies to this author's knowledge, but weed control is still an obvious and necessary recommendation (Lillehoj, 1983). An investigation that compared three cultivation rates to control weeds found no significant differences among the treatments for aflatoxin production in the preharvest crop (Bilgrami et a / . , 1992). The extensiveness of' a weed infestation needed to demonstrate an effect on contamination is, therefore, an academic question that requires no answer in the practical arena.
4. Disease and Insect Involvement Plant disease is normally manifested by unique symptoms and as a reduction in plant vigor. As such, stress on the plant is increased and susceptibility to other organisms is increased, including infection by Aspergillus spp. Numerous diseases are prevalent on corn, all of which have a significant impact on plant vigor, stress, and susceptibility to invasion by fungi such as the Aspergilli. The most critical of these diseases with respect to aflatoxin contamination would be those affecting the ear, especially the ear rots. Although known to be a member of the complex of fungi invading the corn ear, A . f l a w s was not considered to be a seriously damaging ear-rot organism, probably because of its generally nonaggressive nature (Taubenhaus, 1920). Ear rots caused by other organisms such as Helminthosporium have been long associated with the presence of Aspergillus spp. and sometimes with aflatoxin contamination (Doupnik, 1972). Aspergillus
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N. W. WIDSTROM
,flavus is often referred to as an ear-rot organism (Campbell et al., 1993), although now recognized as well for its more notorious reputation as an aflatoxin producer (Campbell and White, 1994). Its presence in the ear-rot complex keeps it available for vigorous activity when conditions favor its development over other organisms. Competition among ear-invading organisms will be discussed in a later section of this chapter. The control of ear rots, stalk rots, and leaf diseases has been accomplished primarily through plant breeding since chemical control is not practical, except when growing specialty corns, sweet corn, or breeding nurseries. Applications of several different fungicides in an experimental situation have been ineffective in significantly reducing aflatoxin contamination (Lillehoj et al., 1984; Duncan et af., 1994). The breeding approach will undoubtedly be necessary in ultimately dealing with the aflatoxin problem. The Aspergilli have long been associated with insect invasion of the corn ear in addition to being members of the fungal ear-rotting complex (Taubenhaus, 1920; Koehler, 1942). The present-day focus on an insect involvement was begun when Anderson et al. (1975) reported preharvest contamination by aflatoxin and its association with insect damage. Sampling studies of harvested and stored corn conducted by the ARS at Peoria, Illinois, also began to show an association of A. j a v u s with insect-damaged corn (Fennel1 er al., 1975, 1977). The association of A. j a v u s and insects was examined in several preharvest field studies (Widstrom et al., 1975; LaPrade and Manwiller, 1977; McMillian et al., 1978; Lillehoj et al., 1978a; Zuber and Lillehoj, 1979) and subsequently the relationships between insects, their damage to ears, and aflatoxin contamination of the corn was clearly demonstrated (Lillehoj et al., 1975b, 1978a). The role of insects in the infection and contamination process has been reviewed extensively (Widstrom, 1979; McMillian, 1983, 1987; Barry 1987). In general, it has been determined that insect damage to the ear is consistently associated with increased sporulations of A . flavus on the ear and increased aflatoxin contamination of the grain (McMillian et ul., 1985b). This concept holds even though other factors may tend to interfere, such as frequent heavy dews that may cause insect damage to increase (McMillian et al., 1985a) and the presence of A. parasiticus that is more closely associated than A . flavus with soil insects (Lillehoj et al., 1980d). Several investigations were initiated to determine which insects were most closely linked to the infection and aflatoxin-producing process (Widstrom et al., 1975). They determined that when confined to ear-feeding, the European corn borer contributed more to the contamination process than either corn earworm or fall arniyworm. The corn earworm is the most frequent lepidopterous ear feeder in the South, and McMillian et al. (1990) found in a 12-year study that A . juvus contamination of the corn earworm moth may also be closely enough associated with preharvest contamination to serve as an early warning system to predict eventual grain contamination by aflatoxin. A series of studies by Guthrie et al.
THE AFLATOXIN PROBLEM WITH CORN GRAIN
243
(1981, 1982) and McMillian et al. (1988) established the European corn borer as a viable contributor to contamination when it occurred as an ear feeder, and only its leaf-feeding habits prevent it from being the dominant insect associated with aflatoxin. The maize weevil (Sirophilus zeamais Motschulsky) is of special interest with respect to the aflatoxin contamination problem because it functions as both a preharvest and storage insect. Initial reports suggested that it was of relatively limited importance and was judged to be a very inefficient vector of A . fravus (LaPrade and Manwiller, 1977). Subsequent studies by McMillian et al. (1980a) demonstrated that the maize weevil can contribute significantly to increased A. jluvus infection on corn ears by transporting spores and damaging kernels. Other investigators later confirmed the maize weevil as being an effective vector of spores and capable of increasing aflatoxin concentration in kernels by as much as 100 times in the presence of the fungus (Rodriguez er al., 1983; Barry et a l . , 1985). Heat and moisture generated by weevil activity in stored corn can be the primary support for A . flavus growth (Dix and All, 1987). A recent addition to the list of vectors is sap beetles (Nitidulidae) that can become carriers of the fungus when both are present in the ear (Lussenhop and Wicklow, 1990). Other ear feeders may also be capable of vectoring the fungus, but are considered unimportant because their frequency as ear invaders is very low. a. Prophylactic Measures There are precautions that can be taken before planting to protect the crop from damage and stress that will occur if insects, diseases, other pests and weeds are not controlled. Such measures are in addition to the buiit-in precautions taken by seed companies that provide protection through seed treatment and inherent resistance of their hybrids to some pests and diseases. The use of prophylactic measures often hinges on the growers’ experience with producing aflatoxin-free corn on the farm, and more specifically in a given field. Cropping history, soil type, availability of irrigation, and experience with site-specific production problems will be determining factors (Smith, 1990). Band application of a nematicide, insecticide, or both at planting is an effective way to protect young plants from early stress and provide a healthy start. This practice is especially important if nematodes or cutworm problems are a part of the field history. Additionally, these treatments are very important for minimum tillage where opportunity for carry-over of problems from the previous year’s crop debris is possible. Many growers also apply pop-up fertilizer at planting to give the corn plant a fast start and improve vigor of the young seedlings. Early control of specific weeds can be achieved by choosing species or weedclass-specific herbicides, preplant incorporated into the top 5-10 cm of soil. Weed problems are dependent on the crop rotation being practiced, weed species
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N. W. WIDSTROM
distribution in the field, effectiveness of the chosen herbicide in controlling weed species that are present, and grower tillage practices. Weed control is a routine recommendation to aid in reducing infection by A . Javus (N. C. Aflatoxin Committee, 1977). In general, weed problems need to be extensive for sufficient stress on the plant to predispose it to contamination by aflatoxin during kernel development (Glover and Krenzer, 1980; Bilgrami et al., 1992). Other sporadic problems can severely affect the level of stress on the young corn plant, such as thrips and nutrient deficiencies due to heavy rains that leach nutrients from the soil. These uncertain occurrences do not warrant prophylactic measures because of economic restrictions. b. Control during the Growing Season It has been a commonly held belief that contamination of the corn crop by aflatoxin is inevitably beyond the control of the producer when conditions conducive to its formation are present. We now know that the risk of A . flavus infection and aflatoxin contamination can be reduced substantially through a good production management system. Events which place young corn plants under stress can have a lasting influence on their susceptibility to attack later in the season. Management toward a healthy crop must begin early. Some stress-inducing events have no remedy and it is already too late for corrective action when they occur (Aldrich et al., 1975). Natural events such as flood or hail are typical examples. Some problems can be diagnosed and corrective action can be taken. Those for which remedial action may be effective require close scrutiny of the crop and immediate action. Insects which attack the very young plant and destroy it completely obviously do not contribute to the aflatoxin problem. Other insects, such as thrips, attack in cool, dry weather and stunt growth of the young plant. An application of irrigation water can often break the infestation cycle and allow plants to recover fully. The most critical time for the growing plant, from the standpoint of aflatoxin contamination, is the grain-filling period (Fig. 1). Any insect attack that produces stress in the plant will increase the risk of Contamination. The best defense against insects in field corn is host plant resistance, a prophylactic tool that is seldom supplemented by insecticides. Insecticides can be effective against leaffeeders and cell-sap feeders, but this method is not often used because of the cost/benefit factor. The greatest impact is made by ear-feeders that feed on kernels and expose damaged tissue for invasion by even the least aggressive fungi, such as the Aspergilli. The concurrence of insect damage and fungal infection of the ear has been recognized for many years (Taubenhaus, 1920). The concurrence concept was extended to harvested corn (Fennel1 et af.,1975; Shotwell er a l . , 1977) and later to preharvest corn (Anderson et al., 1976; Wilson et al., 1981a; McMillian et al., 1985b). Even when fields are checked twice weekly as recommended by extension agronomists (Smith, 1990) there is often little that
THE AFLATOXIN PROBLEM WITH CORN GRAIN
245
can be done to eliminate the infestation after it is established inside the ear. Reducing plant stress by irrigation and timely harvest are other measures that help to minimize aflatoxin contamination. Diseases, like insects, impose stresses on the growing corn plant that render it susceptible to attack by a variety of pests and maladies, including infection by Aspergillus and subsequent aflatoxin contamination. Most disease problems which plague the corn plant after emergence are not easily remedied. Prophylactic measures may be the most effective in controlling disease, especially the use of hybrids that are tolerant or resistant. Those diseases that infect the ear are the most serious contributors to the aflatoxin contamination problem. A complex of ear pathogens and ear-feeding insects all interact (Taubenhaus, 1920) to make effective control difficult. Specific organisms such as Helmitithosporium maydis have been associated as predisposing agents with aflatoxin contamination of the ear (Doupnik, 1972). Certain investigators have chosen to deal with the Aspergilli as ear-rot organisms and evaluate them accordingly (Campbell et al., 1993; Campbell and White, 1994). The Aspergilli are now classified as causative organisms for both ear rots and storage rots by authorities (Shurtleff, 1980), but they were considered by mycologists to be weak or nonaggressive pathogens for many years (Payne, 1987). When diseases do appear, they often occur in localized areas within fields, or only in certain fields (Smith, 1990).Optimum practices in irrigation and fertilization to assure a nonstressed, healthy growing plant can help to minimize the spread of existing diseases and may even limit opportunities for others to get an infection foothold. In essence, management to optimize production is also management to minimize the risk of aflatoxin contamination. Usually by the time disease symptoms are expressed, it is already too late to reverse the process, and emphasis should rather be placed on containment or limiting spread and severity. Any evidence of mold in the ears, found during regular inspections after denting has begun, should initiate grain sampling and testing for the presence of aflatoxin (Smith, 1990). Leaf-feeding by lepidopterous insects during all pretassel stages can impose considerable stress on the plant if damage is extensive. If the infestation is discovered early and is very heavy, an application of insecticide can be effective and economical when one considers potential losses. Decisions regarding insecticide applications on a growing corn crop are very difficult and must be carefully weighed to determine cost benefit, but they must be made quickly, before extensive damage, to be effective. Spot treatment may be effective if infestations are detected early (Aldrich et al., 1975). Experiments designed to determine if aflatoxin contamination could be eliminated if insects were removed from the picture by insecticides were conducted by Lillehoj et a/. ( 1976~)and Widstrom el al. (1976). In both experiments the insecticide treatments did not completely eliminate insect damage nor did they
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N. W. WIDSTROM
preclude A. flavus infection or aflatoxin contamination. Studies by Draughon et a!. (1983) indicated that certain insecticides were capable of inhibiting aflatoxin production by A. parasiticus in the laboratory, and to some extent in the field, but not sufficiently to assure safe use of corn that had been exposed to adequate inoculum. The application of insecticides to control A. flauus infection would certainly not be cost effective unless they could be relied upon to eliminate aflatoxin contamination. Recent studies have confirmed the inadequacy of insecticides as a means to eliminate contamination (Smith and Riley, 1992). Insects such as Heliothis virescens, an insect closely related to some of the earfeeders in corn, are susceptible to aflatoxin (Gudauskas et al., 1967). Aflatoxin also is toxic to several other insects (Matsumura and Knight, 1967), suggesting that A. fluvus or its toxins may function as natural control agents for some insects (Roberts and Yendol, 197 1). McMillian ef al. (1980~)examined this possibility for three of corn’s ear feeders and found that dosages sufficient to adversely affect corn earworm, fall armyworm, and European corn borer ( ~ 2 5 ng 0 g-l) were much higher than allowed as a contaminant of corn as a feed grain (20 ng g-I). The toxicity occurs at such a high concentration that it may be of little practical value. Maize weevils have greater tolerance to aflatoxin than the lepidopterous insects and can survive on grain with contamination levels exceeding 1 Fg g-1 (McMillian et al., 1981). Corn earworm and fall armyworm have lower tolerances to aflatoxin but, as with the maize weevil, they are more drastically affected by A. parasiticus isolates and their toxins than those of A. flavus (Wilson et ul., 1984). Iowa investigators determined that the tolerance of European corn borers increased with each successive instar, and concluded that levels of toxin generated under field conditions might occasionally be great enough to adversely affect the insect, but that the overall influence on the insect population would be minimal (Jarvis et al., 1984). As with diseases, the ultimate insect control method is host plant resistance. The best sources of resistance to various insects will probably be the best option for control, in that neither plant resistance nor insecticides have eliminated damage and that plant resistance is the more cost efficient and environmentally sound. Some resistant germplasm is available for most ear-feeding insects of importance to corn (Guthrie et al., 1970; Guthrie and Dicke, 1972; Scott and Davis, 1981; McMillian et al., 1982b; Widstrom et al., 1983, 1992).
B. MONITORING PROBLEM AREAS Close periodic observations of the corn crop during the early stages of growth and again during the grain-fill period may be critical to minimizing the risk of
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247
eventual aflatoxin contamination. Anything that induces plant stress (moisture deficit, insect infestations, disease incidence, or nutritional deficiency) must be remedied as soon as possible, to prevent the need for lengthy recovery, which provides a wider window of opportunity for vulnerability to attack by Aspergilli. Good management is one of the most important components of producing an aflatoxin-free corn crop, and at worst, a crop with limited contamination in the most stressful environments. With respect to plant stress, those practices that maintain the healthiest highly productive plants also minimize aflatoxin contamination. The growers who maintain good records on crops that were grown in each of their fields and on problems that were encountered during the cropping history of those fields are better able to anticipate problems and take steps to avoid them when corn is again planted in the rotation. A typical example of such records would be a field map showing those areas that are droughty and have produced corn with high levels of aflatoxin in previous years. When droughty areas cannot be avoided, more intensive monitoring of them may serve as an early warning system to determine when conditions are favorable for aflatoxin development. Once these areas are identified, they may be either avoided or eliminated from the harvest when aflatoxin has been detected during years of marginal contamination.
V. HANDLING THE GRAIN CROP AT HARVEST Fortunately, rules for harvest management change very little, whether or not considerations are made for control of aflatoxin Contamination. The basic tenet is to harvest the crop as soon as possible after physiological maturity to maintain grain quality and minimize other losses. The major expense variable at harvest is the consideration concerning artificial drying. This consideration is often a function of weather, especially temperature and moisture, and ultimately the most critical decision to be made for control of aflatoxin once the crop reaches maturity.
A. OPTIMUM TIMING OF HARVEST Early or prompt harvest at maturity is critical in obtaining a crop with minimal aflatoxin Contamination. Delaying the harvest of a crop which is known to have some contamination can only result in higher amounts of aflatoxin in the harvested grain (Jones e? d., 1981). Since contamination is cumulative, delay can only exacerbate the problem on infected ears, even when some resistance to
2 48
N. W. WIDSTROM
contamination is present (Widstrom et al., 1986). Delays of a month or more can cause field losses of 5-10% of the crop in addition to losses due to aflatoxin (Aldrich et ul., 1975). Corn is mature when the grain reaches 300 to 320 g kg-1 moisture (Aldrich et al., 1975). This moisture content is optimal for harvest of grain with the highest yield and quality. With favorable weather, it may be best to wait for some drying in the field when no evidence of A. JQVUS infection is present. When infection is present, and levels of aflatoxin contamination are still less than 20 ng g-I, it may be wise to harvest at a higher moisture and sustain some yield loss to avoid additional contamination of the crop. Harvesting at 260 to 280 g kg-l moisture will minimize harvest losses due to cracking, lodging, shelling, etc. (Lee, 1994). Of these, cracked kernels are the most important contributors to aflatoxin contamination in storage, since they are more susceptible to mold invasion than intact kernels (N. C. Aflatoxin Committee, 1977; Smith, 1981). The infected kernels on an ear are usually widely distributed (Lee et uf., 1980); therefore, isolation or elimination of them will require some modification of harvesting methods (Jones, 1987). Close attention to proper adjustment of the combine is a necessity if cracked kernels are to be avoided and other harvest losses minimized (Glover and Krenzer, 1980). In choosing an appropriate harvest date, the grower must assess the level of contamination present in the field against yield losses due to early harvest at higher moisture and the cost of drying the early harvested crop.
B. G m HANDLINGAND ASSESSMENT OF INFECTION AND CONTAMINATION AT HARVEST An assessment at harvest concerning the presence of A . flavus and level of aflatoxin contamination will to some extent determine the care required in handling the harvested grain. The assessment should be an extension of a scouting program for infected and insect-damaged ears that began when the crop was approaching the dough stage (Duncan, 1979). Meticulous sanitation practices are a must if any evidence of infection or contamination of the preharvest crop is found. Among these are cleaning of auger pits every day, complete cleanout of bins before storage, and maintaining clean trucks, trailers, and combines to prevent further contamination of the crop as it is harvested (N. C. Aflatoxin Committee, 1977; Smith, 1981). Allowing high moisture corn to stand in the wagon for more than 4 to 6 hr is an invitation to rapid build-up of contamination for corn that is already infected with A. fiavus (Glover and Krenzer, 1980). Delays in drying, for whatever reason, can increase contamination 2- to 10-fold (Shotwell et al., 1981, 1983).
THE AFLATOXIN PROBLEM WITH CORN GRAIN
2 49
C. PROCESSING THE CROPFOR MARKET OR HOME STORAGE Drying grain with heated air is necessary when high-moisture corn is harvested since one is at risk for a build-up of aflatoxin contamination unless corn can be reduced to 110 to 130 g kg-l moisture within 48 hr (Lee, 1994). The range in moisture content for safe storage reflects the requirements for different locations. For example, 110 g kg-1 of grain is needed for south Georgia while the cooler climates in the Corn Belt may require only 130 to 140 g kg-' (Aldrich et a / . , 1975). Heat should be used for drying corn with more than 200 g kg-1 moisture in southern states, where aflatoxin contamination is chronic, to avoid spoilage (Smith, 1981). Minimally, sufficient heat should be used to lower the moisture content to 220 g kg-' within 24 hr. Chemical inactivation of the fungi present in the crop can be achieved through the use of acetic and propionic acids when reducing the moisture to acceptable levels cannot be accomplished in the required length of time (Aldrich et ul., 1975). Treatment with these acids does not remove or destroy any aflatoxins that are present, but they merely inactivate or kill the fungus (N. C . Aflatoxin Committee, 1977). The treatments also render the corn unusable for human consumption.
VI. STORAGE AND UTILIZATION OF THE FINAL PRODUCT Vigilance for conditions favorable to infection and increased contamination are vital to storage management and processing in preparation for utilization of corn as food or feed (N. C. Aflatoxin Committee, 1977). The extent of the field infection with Aspergilli will have a profound effect on storability of the crop (Lillehoj and Hesseltine, 1977). Environmental factors quickly impact corn that already is infected (Lillehoj, 1983; Chatterjee rt d., 1990). Moisture is essential for fungal growth (Koehler, 1938), but when temperatures are also high, the combination is ideal for aflatoxin formation (Trenk and Hartman, 1970). Aspergi//usJlavus will grow well and aflatoxin will be produced in high quantities on a variety of substances, including corn (Trucksess et a / ., 1988) when temperatures exceed 30°C and water activities (a,) are greater than 0.86.
A. MONITORING STOREDCORN The heat used to dry the corn grain to 130 g kg-I moisture will not kill Aspergillus spp. The fungi are heat tolerant and can withstand temperatures up to
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46°C indefinitely and as high as 96°C for at least 2 hr (Smith, 1981). Stored corn must, therefore, be inspected on a regular basis for insect activity, moisture, and aflatoxin contamination. Fumigation should be considered if insect problems develop because stored grain-pest infestations consistently contribute to high levels of contamination by aflatoxins (Sinha and Sinha, 1992). If insects become a problem, the moisture level has probably also increased as hot spots have developed, since most insects will not be able to live and reproduce on corn with < 130 g kg-I moisture. Moisture levels may increase throughout the storage area because of high humidity, or hot spots may develop in bins that are not moisture proof against the weather (Shotwell et al., 197%). Hot spots may also develop because of moisture migration within the storage facility. In either case, a regular check for moisture and routine aeration of the corn will be necessary to keep the corn cool and uniformly dry (Smith, 1990; Duncan, 1979). Monitoring for increased levels of aflatoxin contamination is especially important if there have been periods of time when moisture and temperature were above 130 g kg-1 moisture or 30°C. Greater care in sampling must be taken when sampling a storage bin than when sampling from the combine at harvest because the stored lot will be less uniform than freshly harvested corn (Whitaker et a l . , 1979). Samples of a few ears may be adequate for inoculated plants from research plots (Widstrom er al., 1982) while samples of at least 5 kg are considered necessary when probing large seed lots that have not been relocated for extended periods of time (Whitaker et al., 1979). Maintaining a contaminationfree product is as important as having one at harvest, especially if the intent is for use as human food.
B. HUMANCONSUMFTION AND ITS IMPACT ON HEALTH The identification of aflatoxins or causal toxic agents in the sickness and death of animals fed contaminated feed led to a number of international studies to determine if these toxins were related to liver diseases in humans (Shank, 1976), and the extent to which food sources were contaminated. Some studies focused on ingestion and excretion of the compounds (Campbell and Salmut, 1971) and the implications for human health (Campbell and Stoloff, 1974), while others involved the survey of foods and food sources (Jelinek et al., 1989). According to a review of health-related studies, the amount of aflatoxin contamination in food crops and their ingestion regime in Africa and Southeast Asia were related to the incidence of liver cancer (Shank, 1976). The association also was established by several surveys in Thailand and Hong Kong (Shank er al., 1972a,b), in Uganda (Alpert er a / . , 1971), and later in Kenya (Ngindu ef al., 1982). Both animal and human studies revealed aflatoxins as among the most potent hepatocarcinogens known (IARC Monographs, 1987). These toxins are routinely
THE AF'LATOXIN PROBLEM WITH CORN GRAIN
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monitored in body fluids such as urine, blood, breast milk, and serum albumin (Autrup and Autrup, 1992). The liver seems to be the primary target organ, but aflatoxin-contaminated respirable particles in the work environment have also been identified as responsible for an elevated incidence of lung cancer (Hayes et a / . , 1984). The examination of corn dust particles revealed that they can carry sufficient contamination to give cause of concern for the health of workers exposed to the dust (Sorenson et UI.,1981; Burg and Shotwell, 1984; Hill et al., 1984). Some surveys of the food sources for aflatoxin have suggested that the preparatory processes for foods may provide limited protection against ingestion of the toxins (Stoloff, 1979). Initial reports stated that aflatoxin levels were reduced by processing maize beer in Zambia (Lovelace and Nyath, 1977) and in the limewater treatments of maize for producing tortillas (Price and Jorgensen, 1985). Later investigations showed that, though reductions in contamination may be significant, sufficient aflatoxins remained to provide a hazard to human health (Price and Jorgensen, 1985; Okoye, 1986; Carvajal ef a / . , 1987; Arriola et a / . , 1988). Roasting (Conway cr ul., 1978) or hoiling, frying, and baking (Stoloff and Trucksess, 1981; Rehana and Basappa, 1990) were similarly ineffective in eliminating the toxins from contaminated food corn products. A solution to the problem must be found elsewhere.
C. CONTXMINATED CORNAS ANIMAI,FEED A relationship between aflatoxin and toxic hepatitis in farm animals was established by Wilson et a / . (1967). Among economically important farm animals, swine are considered susceptible (LD,, = 0.62 mg kg-I body wt) and chickens relatively resistant (LD,, = 6.5 mg kg-' body wt), while horses, cattle, and sheep fall between these extremes (Pier, 1987). Table 111 illustrates the range of single oral dosages that various species can tolerate. The general effects of aflatoxicosis in several farm animals has not been fully documented (Hintz er al., 1967a,b; Garrett ct al., 1968; Smith and Hamilton, 1970). The laboratory rat was a logical choice of an animal to be used for bioassay of contaminated feeds, but testing showed that the rat was quite resistant to aflatoxicosis effects (Pier, 1987). Test animals were used by some investigators in addition to a chemical assay (Richard and Cysewski, 1971). Although rats have a rather high LD,,. long-term effects are also manifested when feeding contaminated corn with low amounts of aflatoxin is continued (Norred and Morrissey, 1983). Cattle likewise show effects on feed efficiency as well as physiologic, immunologic, and pathologic changes (Richard e t a / ., 1983). Subtle changes that are not immediately apparent are now recognized for most animals. The changes include malabsorption of nutrients, immunosuppression, coagulopathy, poor
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Table 111 Approximate Single-Dose Amounts of Aflatoxin Required for 50% Survival of Several Animal Species Animal Rabbit Duckling Cat Shoat Dog Turkey Guinea pig Sheep Baboon Monkey (cynomologus) Chick Rat Monkey (macaque) Hamster
Single oral LD,, (mg kg-1)
0.30 0.36 0.55 0.62 I .oo I .36 I .40 2.00 2.02 2.20 6.50
5.5- 17.9 7.80 10.20
Source: Pier (1987)
growth, poor feed efficiency, reproductive problems, and increased sensitivity to temperature (Hamilton, 1987; Pier, 1987). These changes also impact the endproducts of meat and milk in different ways. Since aflatoxins contaminate the milk of dairy animals, very low levels of contamination are allowed in their feed (N. C. Aflatoxin Committee, 1978; Smith, 1981). Based on present knowledge, it is now assumed that no level of aflatoxin in feed is free of risk.
D. DECONTAMINATION PROCESSES Considerable effort has been directed toward finding uses for corn that has already been contaminated by aflatoxin. An effective decontamination procedure would be particularly useful in areas of the world where corn is a staple food. The use of heat was an obvious method of choice, but some heating methods reported to remove from 40 to 80% of the contamination (Anderson, 1983) still leave amounts that are unacceptable in foods, i. e., roasting reduces contamination by 80% or less (Conway et al., 1978).Cooking methods removed only 50% of the contamination (Rehana and Basappo, 1990), and several cornmeal preparations for human consumption were even less effective in decontaminating raw corn products (Stoloff and Trucksess, 1981). The liming process in tortilla preparation was also much less effective than originally concluded (Price and Jor-
THE NLATOXIN PROBLEM WITH CORN GRAIN
253
gensen, 1985). Consumption of contaminated corn by humans, therefore, continues to be a serious health problem in third world countries.
1. Blending The blending or dilution process is often used by producers and processors to achieve grade or moisture requirements (Watson, 1987). Christensen and Kaufman ( 1968) noted that this process to achieve either grade or moisture is dangerous and often results in a very high storage risk. The same principle applies to blending corn with different levels of contamination to achieve a level acceptable for feeding, storage, or commercial trade. The practice of blending, illegal in addition to the other risks, is often self-defeating, and is never recommended (Lillehoj and Zuber, 1975). The additional risk of rendering good corn unusable in the process was documented by Lillehoj et af. (1976a) in controlled experiments. Extension circulars have sometimes discussed the process of blending for reduction in aflatoxin contamination (N. C. Aflatoxin Committee, 1977; Smith, 198I), but the benefit of possible salvage of aflatoxin-contaminated grain is greatly outweighed by the numerous reasons that it should not be practiced.
2. Cleaning Mechanical removal of contaminated kernels from seed lots has been largely ineffective. Lillehoj and Wall ( I 987) concluded that using BGYF as an indicator would not work since the identification process requires damaged kernels (Lillehoj e t a / ., 1976d) which are not suitable for electronic sorting. Physical separation or cleaning methods were found to be ineffective for naturally contaminated corn (Brekke et al., 1975a). Anderson ( 1983) suggested that the most optimistic physical methods for separation seem to be associated with density differences between kernels. A significant reduction in contamination was found, for example, in kernels that were nonbuoyant when placed in a sucrose solution (Huff, 1980a,b; Huff and Hagler, 1982). Though impractical as tested, the density principle used may be a useful basis for designing other separation systems.
3. Ammoniation Ammonia has been used for many years as a preservative in controlling the growth of microorganisms in corn (Lancaster et af., 1975; Black et a / . , 1978). Ammonia in a gaseous or aqueous state has been demonstrated as effective in reducing aflatoxin in lots of corn with contamination as high as 1000 ng g-1 to levels below 20 ng g- (Bagley, 1979; Moerek et a / ., 1980; Chakrabarti, I98 1 ). Procedures for ammonia treatment have been developed to handle small or large batches of corn in an economical manner, both in the commercial trade and on
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the farm (Hammond, 1982, 1991). Detoxification by this method is recognized as one of the most promising procedures thus far developed to make utilization of contaminated corn as a feed possible (Anderson, 1983; Lillehoj and Wall, 1987; Bothast, 1991). Several countries and a few states in the U.S.A. have approved ammoniation for decontamination, although the method does not have federal approval in the United States (Phillips et al., 1994). Feeding trials indicate no significant nutritional effects on animals fed the detoxified feed. Studies on feeding of laboratory rats (Norred, 1979, 1982) and farm animals with decontaminated corn indicate no differences in feeding efficiency (Keyl, 1978). The greatest drawbacks in using the ammonia method are associated with the handling of corrosive material and the fact that some animals, swine for example, tend to consume less of the ammoniated corn unless ammonia content of the treated corn is 0.1% or less by dry weight (Jensen er al., 1977).
4. Chemisorbent Additives Another promising method for utilizing aflatoxin-contaminated corn is the application of hydrated sodium calcium aluminosilicate (HSCAS). The compound is a chemisorbent that selectively binds to aflatoxins (Phillips et al., 1991). When added to a contaminated feed diet in amounts of 0.5% or less, HSCAS is able to bind as much as 90% of the aflatoxin content and prevent aflatoxicosis in farm animals (Phillips et al., 1990). The effect of HSCAS was first documented in broiler chicks by Phillips et al. (1988). Similar results were obtained in experiments with swine (Colvin et al., 1989; Harvey et al., 1989). The most recent results demonstrated that HSCAS can prevent aflatoxicosis in turkey poults and greatly decrease the aflatoxin M, residues that occur in milk from lactating dairy cows (Phillips et al., 1991). Further research is needed to determine the specific mode of action by HSCAS, but there is every reason to believe that broad application of this method will provide a means to utilize large quantities of aflatuxin-contaminated corn (Phillips et al., 1994).
E. USEOF CONTAMINATED GRAINFOR ETHANOL PRODUCTION A research effort was directed toward the extended use of the corn plant as a carbohydrate source for ethanol production in the late 1970s and early 1980s, during the “energy crisis’’ (Bagby and Widstrom, 1987). Corn heavily contaminated by aflatoxin was an obvious candidate for use in ethanol production since it could not be used for food or feed. Studies on the fate of aflatoxins in the fermentation process revealed that fermentation did not inactivate aflatoxins or transfer them to the distilled ethanol, but merely concentrated them in the dis-
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tiller’s grain by-products (Dam et al., 1977; Lillehoj, 1978). The spent grains are commonly used as an animal feed supplement and can be an important factor in determining whether conversion of the grain to alcohol is a profitable venture. Decontamination of this by-product is therefore an important consideration for adopting the process of converting contaminated grain to ethanol (Lillehoj et al., 1979). Nofsinger and Bothast (198 1 ) determined that ammonia-detoxified corn could be efficiently converted to ethanol by fermentation with the bacterium Zymomonas mobilis, giving only a slight increase of aflatoxin concentration in the postfermentation solids. Subsequent experiments by Bothast et a / . (1982) revealed that adding ammonia at 1 % of dry weight during liquefaction improved ethanol yields and reduced contamination of the original material by 8 0 4 5 % . A usable supplement for animal feed can be obtained by this process if the level of contamination in the original grain is not excessive (Bothast, 1991).
F. CONTAMINATED GRAINAND
THE
MILLING INDUSTRY
Essentially no effort has been made to utilize contaminated corn in the milling industry, but vigilance is always maintained in an effort to prevent its entry into the industrial processes. A primary reason for the vigilance is that neither drymilling nor wet-milling processes can remove or inactivate aflatoxins (Bothast, 1991). Preventive measures against entry are maintained, although aflatoxins are concentrated in the fraction used for animal feed rather than the food fractions for human consumption from both wet and dry processes, and as such do not pose a great threat to food contamination (Anderson, 1983). Extensive surveys of dry-milling establishments indicate that the screening done by the industry is effective in preventing their use of highly contaminated corn sources (Stoloff and Dalrymple, 1977). The contaminated samples found were primarily from the Southeast and Appalachia. Pilot-plant milling studies have shown that when grain with contamination does enter the system, grits and low-fat fractions always retain the lowest levels of aflatoxin (Brekke er a / ., 1975b; Bennett ef ul., 1976). The wet-milling industry has also been surveyed (Watson and Yahl, 1971), with results similar to those found for dry-milling corn. It is difficult to exclude all contaminated corn because corn purchased by the industry is usually a combination of blends from numerous lots with varied storage history. Most contamination can be traced to high moisture lots. In the wet-milling process, about onethird of the aflatoxin is removed in the steepwater (Yahl ef d.,1971). Similar results obtained by Bennett and Anderson (1978) provide the public with confidence that the potential for significant amounts of aflatoxin-contaminated corn products getting into the human food chain is extremely low.
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VII. LONG-RANGE SOLUTIONS The previous discussion has been focused on stop-gap procedures designed to control the contamination problem until a more permanent solution is available. Most effective permanent solutions require time for development and implementation; thus the stop-gap procedures are vital to adequate interim control measures. The time-proven method of control for pests and diseases is host plant resistance. Initial investigations on the potential of this method for directly controlling Aspergillus spp. infection and aflatoxin contamination were not optimistic (LaPrade and Manwiller, 1976; Zuber, 1977). A pattern of resistance, however, began to emerge, although it appeared to be inextricably related to insect damage and environmental interactions (Lillehoj et a l . , 1976d; LaPrade and Manwiller, 1977). Subsequent screening concluded that evaluation and selection would be difficult, but that the infection and contamination processes were under genetic control (Widstrom et ul., 1978; Zuber and Lillehoj, 1979). A second approach to a long-term solution has been the imposition of chemicals or conditions on the fungus that inhibit aflatoxin production, or the genetic manipulation of fungus or host plant to cause an interruption of the capability of the fungus to produce aflatoxins. The initial efforts in this area involved the use of chemicals (Bothast et al., 1976; Davis and Diener, 1967; Rao and Harein, 1972; Shroeder et a l . , 1974) and atmospheric gases (Landers et a l . , 1967; Wilson and Jay, 1975; Wilson et d . , 1975). Chemicals within the host plant were also suggested as a potential source of aflatoxin production inhibitors (Nagarajan and Bhat, 1972). Another approach to control is the utilization of non-toxin-producing species of Aspergilli or other fungi as competitors that limit growth of toxin-producing strains of the Aspergilli. This approach might arguably be labeled as a stop-gap procedure, since most other invading fungi are also pathogens or contributors to poor grain quality in corn. Calvert et ul. (1978) first compared aflatoxin production of A. flavus or A . parasiticus as competitors and Wicklow et al. (1980) followed with tests of aflatoxin production by Aspergillus spp. in competition with other non-aflatoxin-producing species. Additionally, several isolates of the A. flavus group have been shown to vary in both aflatoxin producing ability and aggressiveness in competing for infection sites (Zummo and Scott, 1994). These findings support the concept of fungal competition as a potential means for control of naturally occurring wild type isolates.
A. BREEDING RESISTANT HYBRIDS The accurate identification of plant types with resistance to invasion by fungi is not possible unless uniform inoculation can be achieved under all field condi-
T H E AF'LATOXIN PROBLEM WITH CORN GRAIN
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tions. The first field inoculation experiments with corn using A. flavus were conducted by Rambo et al. (1974), Anderson et al. (1975), and Lillehoj et al. (1975d). These studies laid the groundwork for more detailed studies to determine the timing and method of inoculation (Calvert et al., 1978; King and Scott, 1982; Widstrom et al., 1981). They concluded that wound inoculation to the ear 20 days after silking was the most effective and efficient method of inoculation to accomplish their purpose of identifying resistant types. Later studies by Tucker et al. (1986), Scott and Zummo (1987), and Scott et al. (1991) gave differing results depending on the objective of the investigation. Wound inoculation proved effective for distinguishing resistant types when only intact kernels were tested for infection percentages. Campbell and White (1994) used massive wounding of the ear with subsequent aflatoxin determination on ear grain samples. Widstrom et al. (1986) demonstrated that waiting until the grain was mature before sampling was the most effective method of demonstrating differences among genotypes. Emphasis was placed on planting adapted rather than resistant hybrids until the 1980s (Glover and Krenzer, 1980). Most reports on screening of genotypes prior to 1980 were lacking in conclusive evidence of the existence of resistance (LaPrade and Manwiller, 1976; Lillehoj et al., 1980b; Manwiller and Fortnum, 1979; Prakash and Siradhana, 1978; Tulpule e t a / ., 1977). A definitive pattern of differences soon began to emerge for resistance to both A . ,fluvus infection and aflatoxin contamination. The differences occurred for popcorns (McMillian et al., 1982a), hybrids (Scott and Zummo, 1990a; Widstrom et al., 1978), and other non-inbred germplasm (Kang et al., 1990; Wallin et al., 1991). Partial explanation for inconsistencies found among genotypes may be found in the fact that factors other than genotypes were the primary objectives under investigation. Some studies were conducted in the laboratory to evaluate testing conditions (Reddy et ul., 1993). others involved inoculation procedures (Wallin, 1986), sampling over time (Lillehoj er al., 1983a), ear traits (Widstrom et al., 1994), and sampling at numerous locations (Lillehoj and Zuber, 1981; Lillehoj et al., 1983b; Scott et al., 1991). Interaction of these factors with genotypes often confuses interpretation of the differences. Zuber (1977) proposed pursuit of genetic differences to improve resistance and followed with genetic investigations (Zuber et al., 1978, 1983) and demonstrated differences among inbred lines using a variety of testing procedures. These and other studies concluded that resistance to infection and contamination were genetically controlled (Darrah et al., 1987; Gardner et al., 1987; Thompson et al., 1984; Widstrom et a/., 1984~).Similar conclusions were drawn when lines other than inbreds were evaluated (Gorman et a / . , 1992; Widstrom et al., 1987). Several important sources of resistance have been identified as a result of recent screening and selection (Campbell e t a / . , 1993; Scott and Zummo, 1988) resulting in the release of resistant germplasm (McMillian et al., 1993; Scott and Zummo, 1990b. 1992). Reviews updating the progress or status of knowledge
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N. W. WIDSTROM
about the genetics of resistance and breeding strategies to develop germplasm have been published from time to time, providing a chronology of technologies that resulted in the development and release of germplasm (Gorman and Kang, 1991; Widstrom and Zuber, 1983; Widstrom et al., 1984b; Widstrom, 1987; Wilson et al., 1989b; Zuber and Lillehoj, 1987). The prospects for finding and developing other sources of resistance are judged as excellent and should, in the future, provide corn with a defense against invasion by A. fravus and subsequent aflatoxin production.
B. INTERRUPTING
TOXIN PRODUCTION BY THE FUNGUS
A basic knowledge of the timing and mode of natural ear infection is needed if effective methods of interruption of the aflatoxin production process are to be developed. The infection process seems to differ considerably when it occurs on kernels of different stages of development. Marsh and Payne (1984a) reported extensive colonization of silks as they dry on the ear and begin to turn brown. The fungus grows along the silks leading to the ear, and eventually penetrates the kernel through cracks in the pericarp and travels through the pedicel areas into the kernel. Zummo (1991) concluded that the primary mode of entry was through the pericarp. The major invasion point of the kernel is apparently the same for the mature kernels in the field and for kernels in storage (Mycock et af., 1988). Formation of aflatoxin occurs earlier under controlled environmental conditions that remain favorable to the fungus than under field conditions (Thompson et al., 1983). The infection process does occur in the field from 1 to 2 weeks after full silk, but aflatoxin accumulation increases throughout the maturation of the kernel and does not peak prior to a kernel moisture of 200 g kg-1 (Payne et al., 1988a). Early infection in both field and controlled environments is probably related to the abundance of simple carbohydrate nutrients in the kernel during the first 2 weeks after inoculation (Widstrom et al. , 1984a). 1. Chemical and Biological Interference in Toxin Production A number of compounds have been tested that inhibit or suppress fungal growth and/or aflatoxin elaboration. Most of them have been evaluated on corn as a stored product and therefore have not been suggested for use as control agents on the living plant. Among these compounds are ferulic acid (Bilgrami et a/., 198lb), potassium metabisulfite (Mabrouk and El-Shayeb, 1981), propionic acid (Shotwell et al., 1984), carbendazim (Devi and Polasa, 1984), 2-chloroethyl phosphoric acid and potassium metabisulfite (Sharma et al., 1987, 1988), sodium bicarbonate (Montville and Goldstein, 1989), several phosphate compounds (Lebron et al., 1989), and several alkenals (Zerinque, 1991). None have been
THE AFLATOXIN PROBLEM WITH CORN GRAIN
2 59
recommended as an economical control on a commercial basis. Biological compounds from sources other than corn, 0-vanillin (Bilgrami et a l . , 1982). neem leaf extracts and volatiles (Bhatnager and McCormick, 1988; Zerinque and Bhatnager, 1994), and spices and their oils (Chatterjee, 1990; Ranjan et al., 1992) have also been shown to influence aflatoxin production in a limited manner. Finally, compounds isolated from corn itself have also been reported as having limited activity against A . j a v u s and aflatoxin production. p-Ionone influenced the morphology and sporulation of Aspergilli (Wilson et al., 1981b) while other volatiles also showed some activity against these fungi (Gueldner et al., 1985). Salt-extracted and base-soluble proteins from corn kernels elicited sporadic fungicidal effects on A . j u v u s but the effects were not always most pronounced in the resistant variety (Neucere and Zerinque, 1987; Neucere and Godshall, 1991; Neucere, 1992). lnconsistent results were also characteristic of evaluations on the activity of chitinase from mature kernels (Neucere et al., 1991). Experiments by Brown et al. (1993) provide evidence that resistance to aflatoxin contamination is related to metabolic activities in the living corn embryo. The mixtures of fungi that occur in nature have been suggested as sources of limitation on growth and development. Ehrlich et al. (1985) demonstrated that aflatoxin production by Aspergillus spp. was lower when grown in mixed cultures with other fungi than when grown as pure cultures. A similar relationship was demonstrated by Horn and Wicklow (1983) and Wicklow e t a / . (1987) when A . j a v u s and A . niger were grown together. The phenomenon is common when A . j a w s is grown with other genera of fungi (Choudhary, 1992; Choudhary and Sinha, 1993; Cuero el a!., 1988; Devi and Polasa, 1987; Wicklow et al., 1988). Fusarium moniliforme is very effective at reducing infection and contamination by A. fravus (Widstrom et ul., 1995; Zummo and Scott, 1992). Aspergillusjuvus and A . parusiticus are both aggressive colonizers of corn, but A . flavus is more persistent on corn than A . parusiticus (Zummo and Scott, 1990, 1994). Brown et a f . (1991) suggested the use of competitive atoxigenic strains of A . j a w s to reduce aflatoxin production on corn. Cotty and Bhatnager (1994) found variation among atoxigenic strains of A . jluvus in their ability to prevent production of aflatoxin and biosynthetic pathway enzymes. None of the chemical and biological methods tested thus far effectively eliminate aflatoxins from contaminated grain; therefore, none have been adopted as fully practical or economical (Bilgrami and Misra, 1981).
2. Manipulation of the Genetics of the Fungus Information about the biosynthesis of aflatoxins slowly began to appear in the literature during the 1970s (Detroy and Hesseltine, 1970; Heathcote et ul., 1976). In the meantime, Papa (1976, 1977, 1979, 1980), working with mutants isolated from A . Jlavus, was developing a linkage map of the fungus, determining
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N. W. WIDSTROM
the inheritance of these mutants, and their interrelationships. Work on the biosynthetic pathway of aflatoxin B, began in earnest with many researchers using A. parasiticus as a fungal model to work out the relationship among intermediates (Bhatnager et al., 1987; Hsieh et al., 1973, 1976a,b; Lee et al., 1976). A continuing development of the body of knowledge on aflatoxin biosynthesis (Bennett and Christensen, 1983) has resulted in new strategies for identifying genes and pathways for different aflatoxins (Bhatnager et a / . , 1989, 1991). An interesting spinoff of work on the genetics of A . parasiticus was the identification of a red-brown mutant that is useful when studying sources of inoculum (Wilson et al., 1986). This mutant produces norsolorinic acid (an orange intermediate) that can be traced in the aleurone layer of corn kernels for the effectiveness of inoculation and tentatively aflatoxin production (Keller er al., 1994). Pending further testing, this mutant may obviate the need for conducting expensive aflatoxin analyses in screening programs. The work with genetics of aflatoxin biosynthesis has already resulted in the cloning of apa-2, a gene associated with regulation of aflatoxin biosynthesis (Chang et al., 1993). New biological control strategies have been proposed based on these findings (Payne, 1992) and a better understanding of the reactions that occur in the biosynthesis of the secondary metabolites, aflatoxins (Bhatnager et al., 1992), brings closer a molecular approach toward understanding aflatoxin production (Keller et al.,1992) and how to control it in preharvest corn.
VIII. CONCLUSIONS Aflatoxin contamination of the corn crop is sporadic in the Corn Belt, but chronic in the South and Southeast. We now have a sufficient body of knowledge concerning the contamination process to establish a framework of guidelines and management practices that will minimize the probability of contamination under a given set of environmental circumstances. Management does not presently provide controls to eliminate, but merely limit or contain contamination by monitoring the crop and following guidelines presented here from preplant through utilization. Fortunately, management to optimize corn production is also management to minimize the risk of aflatoxin contamination. When contamination does occur, options concerning handling and utilization of the crop are available that will help the producer to avoid or at least minimize losses. The advantages of good management in producing, handling, and utilizing the corn crop are obvious, whether or not Contamination by aflatoxin occurs, but longterm solutions are necessary if the problem is to be adequately controlled or resolved. The development of hybrids resistant to infection by Aspergillus spp. and subsequent contamination by aflatoxin is vital in obtaining a lasting solution
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to the problem. In fact, genetic manipulations of both the host plant and the fungus are the most promising avenues of approach for final control and resolution of the aflatoxin problem with corn grain.
ACKNOWLEDGMENTS The author expresses appreciation for helpful discussions with Drs. G. E. Scott and D. M. Wilson, and is grateful for the constructive criticisms during the development of this chapter.
REFERENCES Aldrich. S. R., Scott, W. 0..and Leng, E. R . (1975). “Modern Corn Production,” 2nd ed. A & L Publications, Champaign, IL. Alpert. E., Hutt, M. S. R., Wogan, G . N.. and Davidson, C. S . (1971). Association between allatoxin content of food and hepatonia frequency in Uganda. Cancer 28, 253-260. Anderson, H. W., Nehring. E. W., and Wichser, W. R. (1975). Allatoxin contamination of corn in the tield. Agric. Food Chrm. 23. 775-782. Anderson, R . A . (1983).Detoxification of aflatoxin-contaminated corn. In “Aflatoxin and Asprr~iIIusJlavusin Corn” (U. L. Diener, R. L. Asquith, and J. W. Dickens, Eds.), pp. 87-90. So. Coop. Series Bull. 279. Auburn Univ.. AL. Angle. J. S. (1987). Aflatoxin and aflatoxin-producing fungi in soil. In “Aflatoxin in Maize” (M. S . Zuber, E. B. Lillehoj, and B. L. Renfro, Eds.), Proceedings of the Workshop, pp. 152-163. CIMMYT, Mexico D.F. Arriola, M. C., Porres, E., Cabrerd, S., Zepeda. M.. and Rolz, C. (1988). Aflatoxin fate during alkaline cooking of corn tortilla preparation. J. Agric‘. Food Chrm. 36, 530-533. Aust, S . D.. Albright, J. L . , Olsen, R. E., Byers. J. H . , and Broquist. H. P. (1963). Observations on moldy corn toxicosis. J. Anirn. Sci. 22, 831-832. A u t ~ p ,H . . and Autrup. J. L. (1992). Human exposure to allatoxins-Biological monitoring. In “Handbook of Applied Mycology” (D. Bhatnager, E. 8. Lillehoj, and D. K. Arora, Eds.), Vol. 5, pp. 213-230. Marcel Dekker, New York. and Widstrom. N. W. (1987). Biomass uses and conversions. In “Corn: Chemistry Bagby, M. 0.. and Technology” (S.A. Watson and P. E. Ramstad, Eds). pp. 575-590. Amer. Assoc. Cereal Chem., St. Paul, MN. Bagley. E. B. (1979). Decontamination of corn containing aflatoxin by treatment with ammonia. J. A m . Oil Chem. Soc. 56, 808-8 1 1 . Balzer, I., Bogdanic, C.. and Muzic. S. (1977). Natural contamination of corn (Zeu m a w ) with mycotoxins in Yugoslavia. Ann. Nu/r. Alinr. 31, 425-430. Barry, D. ( I 987). Insects of maize and their association with aflatoxin contamination. I n “Aflatoxin in Maize” (M. S. Zuber. E. B. Lillehoj, and B. L. Renfro, Eds.), Proceedings of the Workshop, pp. 201-21 I . CIMMYT. Mexico. D.F. Barry, D.. Zuber, M. S.,Lillehoj, E. B.. McMillian, W. W., Adams, N . J., Kwolek, W. F.. and Widstroni, N . W. (1985). Evaluation of two arthropod vectors as inoculators of developing maize ears with Aspergillus fkwus. Envirorr. Entonrol. 14, 634-636. Barry. D., Lillehoj. E. B.. Widstrom, N. W., McMillian. W. W.. Zuber, M. S . . Kwolek, W. F., and Guthrie, W. D. (1986).Effect of husk tightness and insect (Lepidoptera) infestation on atlatoxin contamination of preharvest maize. Envimn. G i t o m d . 15. I 116- 1 118.
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N. W. WIDSTROM
Gudauskas, R. T., Homer, E. S., Widstrom, N. W., Thompson, D. L., Bockholt, A. J., and Brewbaker, J. L. (1983). Comparison of open-pollinated maize varieties and hybrids for preharvest aflatoxin contamination in the Southern United States. Planr Dis. 67, 185-187. Zuber, M. S., Darrah, L. L., and Lillehoj, E. B. (1986). International maize genotype-environment study: Natural aflatoxin occurrence. Trop. Sci. 26, 39-43. Zummo, N. (1991). Concurrent infection of individual corn kernels with white and green isolates of Aspergi/!usflavus. Plant Dis.75, 910-913. Zummo, N.,and Scott, G. E. (1990). Relative aggressiveness of Aspergihsflavus and A . parasiricus on maize in Mississippi. Planr Dis. 74, 978-981. Zummo, N.,and Scott, G. E. (1992). Interaction of Fusariurn moniliforme and Aspergillusjavus on kernel infection and aflatoxin contamination in maize ears. Planr Dis. 76, 771-773. Zummo, N., and Scott, G. E. (1994). Pathogenicity of AspergillusJavus group isolates in inoculated maize ears in Mississippi. I n “Biodeterioration Research 4” (G. C. Llewellyn, W.V. Dashek, and C. E. O’Rear, Eds.), pp. 217-224. Plenum Press, New York.
Index A Abortion, pollen and embryo, 205-206 Abscisic acid during dehydrating conditions, 204-205, 210 in xylem sap, 77 Additives, chemisorbent, for contaminated corn, 254 Aflatoxicosis, early identification, 220-22 I Aflatoxins accumulation, affecting factors, 226-236 contamination, and minimizing plant stresses, 236-246 detection and quantification, 221-224 Africa, corn contamination with aflatoxins, 225-226 Aging effect on bioavailability, 70-71 as phytostabilization process, 66 Agricultural research, shifting paradigm, 39-41 Agriculture internal resources, 2 management practices. 4-6 regenerative, 24-26 and soil health, 20-28 Agronomics. in phytoremediation, 94-97 Air contaminants, remediation, 62 quality, soil effects, 15-16 Allernaria macrosporn, as mycoherbicide, I26 Aluminum hydroxy species, balancing, 155- I56 and iron, exchangeable, 158-159 Amendments, organic, see Organic amendments Ammoniation, contaminated corn, 253-254 Animal feeding studies, food produced by different methods, 19-20 Aspergillus flavus. group infection, 226-236 Aspergillus spp., contamination of corn grain, 220-261 Assessments, soil quality and health, 28-44 Atrazine, mineralization rates, 103- 104
281
Augmentation strategy, in weed control, 117 Availability, see also Phytoavailability inorganic P to plants, 145-146 pollutants in soil, 68-71
B Bacteria as plant pathogens in weed control, 124- 129 polychlorinated biphenyl-degrading, 101- 102 Binding, irreversibile, as phytostabilization process, 66 Bioassay, plant enzymatic activity, 63-64 Bioherbicide strategy, in weed control, 116- 117 Biological control aquatic weeds, 123- 124 weeds, seed bank role, 125 Biological interference, toxin production, 258259 BIOMAL, mycoherbicide, 118-1 19 Biomass crops, 193 microbial, turnover, 21 -22 phosphorus, 144 Biotechnological improvements, in phytoremediation, 105- 107 Blending, good corn with contaminated corn, 253 Breeding, atlatoxin-resistant corn hybrids, 256258 Bright greenish-yellow fluorescence test, 223 Bromoxynil, metabolic degradation, 207
C Calcium carbonate, effect on P sorption constants, 165 Canopy, crop, genetic variations. I95 Carbon pools, surface soils, 32-34 CdITotS insecticide uptakes, 78 nitrate content, 18
282 Carrots (continued) remediation of DDT-contaminated soil, 81 Casparian strip, waxy barrier, 74-76 Cato, on choosing a farmstead, 11-12 Cercosporu spp., in control of water hyacinth, 127 Chemical reactions, P in soils, 141-142 Chemicals interference with fungal toxin production, 258-259 synergism with pathogens, in weed control, 128- 129 Chemisorbent additives, for contaminated corn, 254 Chlorinated hydrocarbons remediation in pluntu, 102- 103 rhizosphere degradation, 99- 102 Classical strategy, in weed control, 117, 120123 Clay, adsorption of organics, 70 Cleaning, contaminated corn kernals, 253 Climate, and water use efficiency, 188-192 COLLEGO, control of northern jointvetch, 118 Colletorrichum spp. as microbial pesticide, 118-121 synergy with chemicals, 128- I29 Communities decomposer, 8 soil fauna, 20-24 Complexation reactions, organic acids with metals, 147-148 Composting rock phosphate solubilized during, 167 for soil health, 14 in tightening nutrient loop, 24 Contamination coin with aflatoxin and hybrid selection, 231 -236 and minimizing plant stresses, 236-246 scope, 224-226 groundwater, 15-16, 26-27 heavy metal and toxic element, 17 and infection, assessment at harvest, 248 soils, phytoremediation, 55- 107 Coin aflatoxins as contaminants, 220-226 contaminated, as animal feed, 251-252 grain crop, handling at harvest, 247-249 growth and ear development, 236-247 preplanting considerations, 229-236 stored, monitoring, 249-250
costs economic, of remediation, 60-61 environmental, of agricultural production, 26-28 phytoremediation, 66-67 Cmsulacean acid metabolism, and water saving, 191 Cropping systems advocated by philosophers, 11- 12 legume-based, 26 Crop species microsporogenesis during drought, 204-205 nutritional needs, 18 in phytoremediation agronomics, 94-97 water use efficiency, 189- 192 Crop yield indicator of sustainability, 44 nondeclining trends, 24 Crude oil, phytotoxic agents in, 98 Cultivation, and aflatoxin contamination, 24024 I Cultural heritage, soil as, 8 Cuticle, importance to drought tolerance, 203204
D DDT, contaminated soil, remediation with carrots, 81 Decomposition processes, soil, 8 Decontamination, see also Phytodecontamination processes aflatoxin-contaminated corn, 252-254 pesticides, with trees, 62 remediation technique, 57-58 Degradation environmental, 41-42 in rhizosphere, 87-88, 99-104 soil, 5 Dehydration postponement and tolerance, 196- 197 seeds, 207-210 Desiccation tolerance, molecular features, 208210 Desorption, phosphorus under flooding-aerobic cycles, 175- 176 organic amendment effects, 163- 165 Disease resistance of corn hybrids, 233-236 role in atlatoxin contamination, 241-246 Distribution, pollutants in soil, 67-68
INDEX Diversity, in crop production, 25 Drinking water, presence of nitrate, 15- 16 Drought tolerance, improvements, 197-204 Dry mass, and water use efficiency, 189-192 Dry matter, relationship to water use, 198-199
E Ear development, managing conditions during, 236-247 Earthwornis, indicative of healthy soil, 22 Ecological function$, soil, 8- 10 Economics, of remediation, 60-61 Ecosystem, soil as, 6-8 Edaphic factors, affecting atlatoxin accumulation, 228-229 Eh, see Oxidation-reduction system Embryo, abortion due to water deficit, 205-206 Embryo maturation proteins, 209-210 Endophytes. commercial use, 88-89 Environmental costs, agricultural production, 26-28 Environmental impact, microbial herbicides, 129- 13I Environmental quality, and sustainability, 3-6 Enzymes, desiccation effects, 208-209 Equation of life, 7-8 Ethanol production, use of contaminated grain, 254-255 Evapotranspiration, crop, 191- 193
F Farming, organic versu.y conventional, 18-20. 32-34 Fate contaminant, in soil-microbe-plant system. 63 phosphorus, from organic amendments, 160I63 xenobiotic, in plant system, 74-82 Feed aflatoxins in, detection methods, 222 animal, contaminated corn as, 25 1-252 Fermentation, atlatoxin-contaminated corn, 254255 Fertility, Soil, maintenance, 12 Fertilization, relationship to aflatoxin contamination, 239-240 Flooding-aerobic cycles, and P desorption, 175- I76
283
Food quality, soil effects, 16-20 Food web, microbial herbicide effects, 130- I3 I Forests, and global C balance, 37 Fulvic acid, effect on P sorption, 148-154 Fungus, see also speciJicfungi chemical inactivation, 249 ectomyconhizal, in bioremediation, 101 toxin production, interruption, 258-260 in weed control, 117-132 white-rot, targeted to pollutants, 84
G Gaia hypothesis, 23 Gas exchange efficiency, compared to water use efficiency, 193- 194 Genes, degradative, 105- 106 Genetic manipulation, fungus, 259-260 Genetic selection for earliness, 200-202 in enhancing phytoremediation, 95-96 Germplasm, corn, non-inbred, 257-258 Global function, and sustainability, 3-6 Grain contaminated for ethanol production, 254-255 and milling industry, 255 corn, adatoxin problem, 219-261 Grain-filling period, and aflatoxin contamination, 244-245 Grasses native range, genetic improvement, 203-204 as soil stabilizers, 99 Groundsel, control with rusts, 122- 123 Groundwater bioremediation, 56 contamination with agrochemicals, 26-27 nitrate-contaminated, 15- 16 Growing season, aflatoxin control during, 244246
H Harvest, corn, optimum timing, 247-248 Harvest index, part of water use efficiency, 192I93 Hazardous waste, definition, 93-94 Health human, impact of corn adatoxins, 250-251 human and animal, soil health effects, 14-20 rangeland, 37
2 84
INDEX
Health (continued) soil definition, 10- I 1 and sustainability, 1-45 Henry's Law, 72 Herbicides, see also Mycoherbicides microbes as, environmental impact, 129- 131 selectivity due to plant metabolism, 80-81 weed class-specific, 243-244 Horseradish, potential for soil remediation, 8384 Humic acid, effect on P sorption, 148-154 Humification, as phytostahilization process, 65, 83-84 Humus prevention of phosphate fixation, 152- 153 soil quality indicator, 13 Husk tightness in corn hybrid selection, 23 1-232 and resistance to insects, 234-235 Hybrids corn breeding, 256-258 selection for aflatoxin resistance, 23 1-236 maize, high dry-matter approach, 199 Hydrated sodium calcium aluminosilicate, in corn decontamination, 254 Hydrocarbons, total petroleum, cleanup, 69 Hydrogen peroxide, in phytoremediation, 84
I Immobilization, and mineralization, P, 143144, 160-163 Imrnunochemical methods, in aflatoxin detection, 223-224 Indicators key, threshold values, 34-35 soil health earthworms, 22 use of minimum data set, 29-34 Infection Aspergillus flaws, 226-236 effect of maize weevil, 243, 246 and contamination, assessment at harvest, 248 mycorrhizal, in plant survival, 87 lnositol phosphate esters, in soil, 143-144 Insecticides plant-produced, 78-79 use on growing corn crop, 245-246
Insects resistance of corn hybrids, 233-236 role in aflatoxin contamination, 231, 241-246 synergism with pathogens in weed control, 126- I27 International conferences, soil sustainability, 34 Intrinsic value concept, 9 Invertebrates, classification, 21-22 Invert emulsions, in weed control with hiological pesticides, 128-129 Iron and aluminum, exchangeable, 158- 159 crystalline ferrous hydroxides, 173- 174 Fe (111)-bound phosphate, 170-171 Irrigation, corn plants, 236-239 Isotope ratio, plant tissue, 194-195
K KO,
lipophilicity related to, 73 in phytoremediation models, 92 transpiration stream concentration factor related to, 75-76
1 Land use, in determination of exposure to contaminant, 59 Leaching, soluble compounds during phytoremediation, 98 Lignification, as phytostabilization process, 66 Lignin, prevention of phosphate fixation, 152I53 Lipophilic compounds, associated with soil surfaces, 69-70 Lipophilicity, related to KO,, 73 Liver, plant analogy, 80 Losses, reproductive, related to water deficits, 204-207, 21 I
M Maize cell enlargement, 204 hybrid, high dry-matter approach, 199 Maize weevil, effect on Aspergillusflavus infection, 243, 246
INDEX Management practices agricultural, 4-6 during corn growth and ear development, 236-247 farm, integration of soil health concepts, 39-
44 and food quality, 18-20 optimizing corn production, 260 Manure animal, and plant residues in aerobic soils, 156-167 effect on phosphorus availability, 145-146 Metabolism, xenobiotics in plants, 78-8 I Metarhizium anisopliue. effects on nontarget species, 131 Microbes biomass, turnover, 21 -22 as herbicides, environmental impact, 129- 13 1 populations in bulk soil, 86-88 in weed control, synergisms affecting, 125-129 Microflora, plant-associated, in remediation, 85-90 Milling industry, and contaminated grain, 255 Mineralization atrazine, 103-104 chlorinated hydrocarbons, 100-101 and immobilization, P, 143- 144, 160- 163 organic P in flooded soils, 17 I - 172 Minimum data set. in assessment of soil health, 29-34 Modeling phytoremediation, 91 -92 speciation, 159 Models, surface adsorption, 142 Moisture levels in stored corn. 250 soil, monitoring, 238 Molecular biology, and water use efficiency, 195-196 Monitoring corn, during growth and grain-till, 246-247 nutrients, during phytoremediation, 96-97 soil moisture. 238 stored corn, 249-250 Mutants, isolated from AsperRil/usJhus, 259260 Mycoherbicides, see also Herbicides Alrernaria macrospora, I26 BIOMAL, 118-1 I9 Mycorrhizal infection, in plant survival, 87
285
N Natural resource accounting, 26-28 Nematicides, band application before planting corn, 243 New Zealand, biodynamic and conventional farming, 32-34 Nitrate In drinking water, 15-16 free, antinutritive factor in food plants, 17-18 Nitrogen effect on aflatoxin contamination. 239 mineralizable, 32-34 Nitrosamines, nitrate conversion to, 15- 17 Northern jointvetch, control COLLEGO, I 18 pathogenlinsect synergy, I27 Nutrient cycling, 6-9, 23-2.5 Nutrients monitored during phytoremediation, 96-97 replacement costs, 27 Nutrition, plant, 13-14, 23, 239-240
0 Organic acids effects on surface charge, 155- 156 role in P solubilization, 154-155 in soils, 146-147 complexation with metals, 146- 148 Organic amendments effects P reactions, 144-146 P sorption, I7 I - I76 and Eh and pH, in waterlogged soils, 167-170 enriched with inorganic P, 165-167 P fate from, 160-163 Organic matter, soil, see Soil organic matter Organic production, comparison with conventional farming, 18-20, 32-34 Organics plants as remediation structure, 71-82 soils contaminated with, remediation, 56-61 Orthophosphate adsorption, 155-156 and oxalate, sorbing site competition, 150-
I51 reaction with soil constituents, 144- 146 Osmotic adjustment during water deficit, 197-198 wheat and cotton, 200
286
INDEX
Ovary starch, mobilizable, 206 Oxalate, competition with P for sorbing sites, 150- 15 1 Oxidation-reduction system, and organic amendments, 167-168 Oxygen, partial pressure decrease, 97-98
P Pathogens, fungal and bacterial, in weed control, 116-129 Pentachlorophenol, mineralization, LOO Pesticides assessment for health effects, 16-17 microbial, in weed control, 115- I32 plant uptake, 73 rhizosphere degradation, 103- 104 Petroleum contamination, phytoremediation, 99 Petroleum hydrocarbons, total, cleanup, 69 PH and organic amendments, in waterlogged soils, 168-170 soil adjustments, 230 organic residue effects, 156- 158 as soil quality attribute, 34 Philosophers, Roman, proponents of soil health, 11-12 Phosphate phytoavailability, 156 precipitated, dissolution, 154- 155
rock solubilization, 167 utilization, 145 Phosphoric acid, solubility, 145 Phosphorus content of organic residues, 160 cycle, in soils, 140-144 inorganic, organic amendments enriched with, 165-167 from organic amendments, fate, 160-163 solubility in flooded soils, 170- I71 sorption organic amendment effects, I7 I - 176 plant residue effects, 163-165 sorption sites, competition, 148-154 Phyllosphere, microbiota composition, 89-90 Physicochemical effects, plant-produced, 90-91 Physiology, water use efficiency, 189- 192
Phytoavailability phosphate, 156 phosphorus, 165-167 Phytodecontamination processes, 64 Phytoextraction, important soil organic contaminants, 76-78 Phytophthora. in weed control, 117-1 18 Phytoremediation apparent advantages, 66-67 biotechnological improvements, 105- 107 concepts and definitions, 61-66 ex planfa. 82-91 limitations, 97-98 modeling, 91-92 site conditions and limitations, 92-94 Phytostabilization acceptability, 69-70 processes, 64-66 Planting date, associated aflatoxin contamination, 233 Plant pathogens, fungal and bacterial, in weed control, 116-129 Plant population, and risk of aflatoxin contamination, 232 Plants associated microflora, 85-90 C3 and C4, 95, 190-191, 196 drought tolcrance, 187-21 I enzymatic effects ex planfa, 82-85 green redefinition, 66 in remediation, 61-66 nontarget, microbial herbicide effects, 129- 13I nutrition, 13-14, 23, 239-240 physicochemical effects, 90-9 1 recombinant field trials, 67 for phytoremediation, 105-106 as remediation structure for organics, 71-82 transformation, 106 Pliny, on earth musty odor, 12 Poisoning, with contaminated soil, 15 Pollen abortion, 205 desiccation, 208 shed, early silking relative to, 206 Pollutants distribution and availability, 67-71 organic, contaminated soil, remediation, 56-61
INDEX Polychlorinated biphenyls, bacteria degrading, 101-102 Polycyclic aromatic hydrocarbons contaminated soil, 99 volatilization, 72 Pools carbon, 32-34 phosphorus, 140-141 Precipitation, phosphate, 154- 155 Preharvest, aflatoxin contamination, 221 -222, 224-225, 234, 242-243 Preplanting, considerations for aflatoxin accumulation, 229-236 Prickly pear, control with pathogen/insect synergy, 126- 127 Producers. interaction with scientists, 39-44 Prophylactic measures, before planting corn, 243-244 Proteins, embryo maturation, 209-210 Purcinia spp., evaluation for weed control in U . S . . 120-123
R Recombinant plants field trials, 67 in phytoremediation, 105- 106 Record of decision, in cleanup process, 60 Regionalization, aflatoxin contamination, 227 Regulatory extraction protocols, 7 I , 83 Remediation, see also Phytoremediation in planro, 102-103 plant-associated microflora in, 85-90 soils conrdminated with organics, 56-58 economics, 60-61 process and goals, 58-60 Reproduction, and water deficits, 204-207 Research agricultural, shifting paradigm, 39-41 P sorption and role of organic amendments, 176-178 Residues, organic effect on pH, 156-158 and exchangeable Al and Fe. 158-159 P content, 160 Resistance corn hybrids to insects and disease, 233-236 host plant, hybrid breeding, 256-258
287
Resources finite, soil as, 3-1 I internal, agriculture, 2 natural, accounting, 26-28 Respiration, soil, farming method effects, 3234 Revegetation, spontaneous, with hardy weed species, 79-80 Rhizosphere biasing, 85, 106-107 degradation with chlorinated hydrocarbons, 99- I02 pesticides, 103-104 structure and function, 86-87 Rice genetic variation for rooting, 202 upland and paddy, 196 Risk, in soil remediation, 59 Rock phosphate solubilization, 167 utilization, 145 Rodale, J. I . , 13-14 Rodale. Robert, 24-26 Root concentration factor, 75 Root exudate, as nutrient for microorganisms, 86-88 Rooting, deep, importance to drought tolerance, 202-203 Roots peroxidase activities, 84 redefinition, 66 uptake of xenobiotics, 71-74 Rotations based on plant progressions, 25 crop, associated soil microflora, 230-231 Rush skeletonweed control with Pucciniu spp., 121 infestations, 116 Rust, evaluation for weed control in U.S., 120123
S Sandy soils, water-holding capacity, 238 Scientists interaction with producers, 39-44 19th and 20th century, on soil vitality, 13-14 Sclerotinia .demtiorurn. control of thistle. 1 19I20
288 Seed banks, microbial management, in weed control, 125 Seeds, dehydration, 207-210 Selenium, soil concentrations, 17 Septoria passiflorae, control of Passifora weeds, 121- 122 Sequestration, xenobiotics within plants, 81-82 Simazine, uptake by barley, 74-75 Sites, remediable, conditions and limitations, 92-94 Soil aerobic, 146-167 attributes, estimation, 37-39 bulk, microbial populations, 86-88 contaminated with organic pollutants, remediation, 56-61 cultivated, aflatoxin spore load, 228 faunal communities, 20-24 as finite resource, 3- I 1 flooded, and P solubility, 170- I7 1 moisture monitoring, 238 P cycle, 140-144 PH adjustments, 230 organic residue effects, 156- 158 phases, 72-74 pollutants, distribution and availability, 6771 waterlogged, and P behavior, 167-176 Soil amendments, organic, see Organic amendments Soil bulk density, and use of soil indicator ratios, 32-33 Soil depreciation allowance, 27-28 Soil health and agriculture, 20-28 assessment, 28-44 comparison with soil quality, 6- 1I concepts early proponents, 1 1 - 14 integration into farm management, 41-44 and human health, 14-20 Soil organic matter effects P fixation in flooded soils, 172-176 surface charge, 155-156 replenishment, 24-26 Soil quality assessment, 28-44 index, 36-37
comparison with soil health, 6-1 I test kit, 43-44 Soil water. use by sorghum genotypes, 200-201 Solid phase, soil, 74 Sorghum, genotypes, usage of soil water, 20020 1 Sotption, phosphorus humic and fulvic acid effects, 148-154 organic amendment effects, 171-176 plant residue effects, 163- 165 in soils, 141-142 Speciation modeling, 159 Spontaneous revegetation, with hardy weed species, 79-80 Spurge, control with rust fungi, 122 Stabilization, remediation technique, 57-58 Starch, ovary, mobilizable, 206 Storage, and utilization of final corn crop product, 249-255 Stress, corn plant, minimization, 236-246 Sugar replacement hypothesis, 209 Suitability criteria, met by soil health indicators, 30 Surface charge, organic acid effects, 155-156 Survival, plant, role of mycorrhizal infection, 87 Sustainability and global function, 3-6 and soil health, 1-45 Synergism, pathogens with chemicals, 128- I29 insects, 126- 127 other pathogens, 125- 126
T Technology transfer, in soil assessment, 4244 Temperature associated with aflatoxin accumulation, 227228 leaf, effect on vapor pressure, 189 soil, and planting time, 95 Test kit, soil quality, 43-44 Thin-layer chromatography, allatoxins, 223 Thistle, control with fungi and rusts, 119-122 Threshold values, for key indicators, 34-35 Tillage and aflatoxin contamination, 240-241 minimization, 25 Tissue sampling, for nutritional problems, 240
2 89
INDEX Tolerance desiccation, molecular features, 208-210 drought, improvements, 197-204 Tomatoes organically and conventionally managed, 25 water use efficiency, 193-195 Total petroleum hydrocarbons, cleanup, 69 Toxin production, by fungus, interruption, 258259 Trait selection for earliness, 200-202 in enhancing phytoremediation, 95-96 Transformation phosphorus, 142-144, 160-163 plant, 106 Transpiration stream concentration factor, 7476 Transport, xenobiotics in plants, 74-78 Trees, in pesticide decontamination, 62 Trichloroethylene mineralization, 100- 101 plant uptake, 102-103 Trinitrotoluene, phytotoxicity, 98
U United States corn contamination with aflatoxins, 224-225 weed control with fuccinia spp., 120-123
quality, soil effects, 15-16 saving, and crassulacean acid metabolism, 191 soil, use by sorghum genotypes, 200-201 Water hyacinth, control with Cercospora spp., 127 Water phase, soil, 73 Water use, normalization for evaporative demand, 191-192 Water use efficiency definition and physiology, 188- 192 measurement, 193-195 and molecular biology, 195- 196 unchanged with stable water supply, 21 1 Wax deposition, native range grasses, 203-204 Weather, effect on aflatoxin accumulation, 226228 Weeds aquatic, biological control, 123- 124 control and aflatoxin contamination, 240-241 with plant pathogens, 116-124 as volunteers in contaminated soils, 79-80 Wetlands, constructed and natural, 61-62 Wheat genetic selection, 201 -202 osmotic adjustment, 200 Wound inoculation, corn ear, 233-234, 257
X
V Vapor phase, soil, 72 Varro, on farm soil quality, 11-12 Virgil, on maintaining soil fertility, 12 Vitality, soil, 13-14
w Waste water, municipal, contaminant removal, 62 Water, see also Groundwater deficits, and reproduction, 204-207 drinking, presence of nitrate, 15-16 holding capacity of sandy soils, 238
Xenobiotics fate in plant system, 74-82 in soil, 67-71 uptake into roots, 71-74 Xylem sap, obtaining and measuring, 76-77
Y Yield crop indicator of sustainability, 44 nondeclining trends, 24 relationship to drought tolerance, 199-200
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